Published in 2014 in the Journal of Environmental Management 135: 100-09
The use of replacement cost method to assess and manage the impacts
of water resource development on Australian indigenous customary
economies.
Sue Jackson1 , Marcus Finn2 and Kelly Scheepers3
1 Corresponding author
Principal Research Scientist, CSIRO Ecosystem Sciences, PMB 44, Winnellie, NT 0822, Australia.
Present address: Australian Rivers Institute, Griffith University, Nathan, Queensland 4111, Australia.
Telephone: +61 3 94865581
Email: [email protected]
2 Research Scientist, CSIRO Ecosystem Sciences, PMB 44, Winnellie NT 0822, Australia
Present address: Environmental Consultant, Farrer, ACT 2607 , Australia [email protected]
3 Postdoctoral fellow, CSIRO Ecosystem Sciences, PMB 44, Winnellie, NT 0822 ,Australia. Email: [email protected]
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Abstract
The value of functional and biologically diverse landscapes to
indigenous societies is increasingly recognised in public debates
about development pathways but rarely rigorously assessed in
development decisions. Using the replacement cost method, we
quantify the direct consumptive value of aquatic species and sites
for indigenous subsistence in three Australian tropical river
catchments where negligible data exists on indigenous water values
and the extensive use of wild resources for food, art, craft and
medicines. The results establish a baseline for assessing and
monitoring the socio-economic impact of hydrological and
ecological changes from water resource development. More than 90%
of the gross replacement value in each catchment was accounted for
by a small subset of high value species which could be used as
integrated indicators of ecological and socio-economic change. The
total value of species harvested was distributed across a large
number of sites, justifying the need for a regional management
approach to ensure the maintenance of diverse habitats for hunting
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Published in 2014 in the Journal of Environmental Management 135: 100-09
and fishing. While ‘value’ is a cultural, context-dependent
construct, studies like this one can lend legitimacy to a targeted
approach to environmental and social impact assessment of water
resource development proposals by calling for prioritisation of
mitigation and management actions.
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1.0 Introduction
It is widely accepted that freshwater ecosystems face multiple and
severe threats rendering them one of the world’s most imperilled
ecosystems. Significant environmental damage has occurred as a
result of river regulation and alteration of flow, catchment
disturbance and habitat loss; pressures that are driven by the
escalating human demand for land and water concomitant with
widespread industrial transformation (Arthington et al., 2010;
Millennium Ecosystem Assessment, 2005). In addition to the many
ecological impacts, degradation of the world’s freshwater
resources has profound socio-economic consequences affecting the
entire human population (Vorosmarty et al., 2010). Human
interactions with water bodies and aquatic ecosystems are a source
of cultural inspiration and often have a religious foundation.
Freshwater ecosystems underpin global food production based on
commercial and subsistence fisheries, aquaculture, agriculture and
pastoral grazing (Arthington et al., 2010). Many indigenous
collectivities rely on aquatic resources for their livelihoods, and
it is these groups that are most vulnerable to the impact of water
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resource development projects, as their interests are often
overlooked and they stand to benefit least from river development
(King and Brown , 2010; see also Thomas and Twyman, 2006).
Northern Australia’s rivers and wetlands are environments of vital
importance to the numerous indigenous language groups who are
closely connected to these landscapes via customary tenure and use
of aquatic resources, cosmological beliefs and a body of
environmental knowledge accumulated over hundreds of generations.
Almost a third of northern Australia’s 300,000 inhabitants are
indigenous and, as a result of post-colonial land claims, now own
approximately a quarter of the region’s land mass (Altman and
Jackson, 2009). The strength of customary relationships with rivers
and aquatic ecosystems is now formally recognised in national
native title law, national water policy and in efforts to address
the ‘cultural values’ of water in regional natural resource
management strategies and development decisions.
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Indigenous knowledge and values therefore need to be reflected in
environmental assessments and management actions undertaken in
northern Australia, whether environmental impact studies (EISs)
prepared for agricultural or mineral development impacting on water
sources or environmental flow assessments (EFAs) conducted as part
of water allocation planning exercises. This is particularly so for
the region under study because there is considerable contestation
over development visions which manifest in the popular and
recurrent call to increase agricultural exports (Hamilton and
Gehrke, 2005).
With 55 river catchments covering an area of more than 1.3 million
km2 (Stoeckl et al., 2006), Australia’s n orth represents the world’s
largest intact tropical savannah (Woinarski et al., 2007) and
contains one of its highest concentrations of free-flowing systems
(Australian Tropical Rivers Group, 2004). The region comprises 15%
of Australia’s landmass yet its rivers account for about 65% of the
country’s runoff. S peculation about the potential to harness this
resource has been constant over a recent decadal drought as the
volume of water stored in major reservoirs in southern Australian6
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declined (Blanch, 2008) . The Federal Government magnified the
focus on the potentially valuable northern water resource in 2007
when it released a National Plan for Water Security. This plan, inter
alia , established the North Australia Land and Water Taskforce with
the aim of examining the potential for further land and water
development, particularly irrigated agricultural development. A
key recommendation of that inquiry highlights an essential research
priority for the region:
Australian governments should significantly increase investment in social,
cultural, and economic analysis to support the assessment of competing values
and uses for land and water use planning, catchment level water planning and
local decision making (2009:3).
Quantifying ecological goods and services provided by rivers and
establishing better links between environmental flows, organism
health and human well-being are therefore important research
priorities (Stoeckl et al., 2013; Zander and Straton, 2010). In this
region, lack of knowledge of ecosystem patterns, processes and
response to anthropogenic impacts represents a significant
constraint on development planning, water allocation planning and7
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statutory environmental impact assessment, as does insufficient
understanding of the socio-economic consequences. Baseline
information with which to assess changes in ecological condition is
inadequate for most Australian tropical river systems (Douglas et
al., 2011) and a lack of quantitative information to assess the
volume of water that could potentially be extracted has hampered
rational d ebate about irrigation development ( Petheram et al.,
2010) . In addition, the contribution that aquatic ecosystems make
to economic and social well-being is not well understood (Stoeckl et
al., 2006).
For environmental assessments to adequately protect the interests
of indigenous communities, they require a sound understanding of
the socio-economic and cultural significance of the environment,
particularly the degree of reliance on natural resources. Yet
Altman et al. (2012) argue that the benefits of customary or
subsistence harvest undertaken by indigenous people are
undervalued by Australian society. The relatively little
quantitative research on indigenous use of natural resources
appears to reflect that undervaluation (Altman, 1987; Asafu-8
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Adjaye, 1996; Gray and Altman, 2006; Venn and Quiggin, 2007; Walsh
and Douglas, 2011). The ‘hidden’ nature of the value of natural
resources to indigenous households provides challenges for
economic valuation methods that rely on functioning markets, prices
and demand driven measures (Field, 2002).
As one approach for assigning monetary values to a part of the larger
total economic value (TEV) of these non-market ecosystem goods and
service to indigenous households, this study uses the replacement
cost method. The replacement cost method looks at the cost of
replacing a damaged or lost asset and uses this cost as partial proxy
or measure of its value . The replacement cost method is not a demand-
based approach, and cannot be expected to generate valid or "true"
welfare measures (i.e., Marshallian or Hicksian welfare change
measures) according to economic theory, and may reflect only a part
of the larger TEV. However, the approach may still be a useful
heuristic tool in any project appraisal, and provide valuable
information to policy makers (see Turner et al. , 1993).
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The aim of this paper is to demonstrate the applicability of the
replacement cost method to assess and manage the impacts of water
resource development on the customary economies of indigenous
societies. Drawing on results from two studies of indigenous
subsistence patterns in three tropical Australian catchments (see
Jackson, 2011; Scheepers, 2012) we value the goods and provisioning
services generated by tropical rivers that are of benefit to
indigenous people for food, fibre, and medicines. We show that the
replacement cost method is a useful way of determining the species
of greatest economic value to subsistence resource users, aiding
environmental flow risk assessments and improving environmental
management.
Valuation calculations are most commonly used to feed into economic
or investment appraisal processes and a pplications to
environmental assessments remain low in number (Baker et al.,
2013). Understanding the relative importance of particular
components of the ecosystem provides critical information upon
which environmental standards can be set to maintain socially
acceptable ecological conditions (Finn and Jackson, 2011; Poff et10
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al., 2010). The results can serve as a baseline for long term
monitoring of indigenous water values.
The paper is structured as follows. We first describe our study
sites, briefly explaining the management context affecting land and
water use in these sites and then outline methods of data
collection. The paper will then turn to the results and discussion.
2.0 Data and method
2.1 The study area
Our study area comprises three large river systems in north
Australia: the Daly (Northern Territory), Fitzroy (Western
Australia) and Mitchell (Queensland) (Figure 1). North Australia
is mostly flat and consists of extensive, low, inward draining
plateaus, which are situated behind narrow coastal plains (Petheram
et al., 2008). More than 94% of the region’s rainfall is received
between November and April, generated by local convection and
tropical cyclones (CSIRO, 2009). When the region’s hydrological
characteristics are compared with the rest of the world,11
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variability and seasonality of annual stream flows are high
(Petheram et al., 2008).
Insert Figure 1 around here
All three rivers display distinct hydrological seasonality with
most of their discharge occurring predominantly during the wet
season (November to April), while inter-annual variation in the
magnitude and timing of peak flows is high (Kennard et al., 2010).
During the dry season (May to October), perennial flow in the main
channels of the Daly and Mitchell Rivers is sustained by groundwater
inputs, whereas the Fitzroy River’s main channel ceases to flow and
is classified as highly to extremely intermittent (Pusey et al.,
2009). In this environment, permanent pools fed by groundwater
provide crucial ecological refuges during exceptionally hot, dry
conditions.
The Fitzroy River catchment has a population of approximately 7,000
people, of whom 64% are indigenous (Larson and Alexandridis, 2009);
the Daly River’s population is 10,000 people of whom 28% are12
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indigenous (Larson and Alexandridis, 2009); and t he population of
the Mitchell River catchment is just over 5,500, of whom 23% are
indigenous (Larson and Alexandridis, 2009). Indigenous socio-
economic status tends to be very low: the case study regions were
described as ‘regions of disadvantage’ by Biddle (2009). The number
of indigenous residents per household is relatively high: in 2006
there were 3.8 people per household in Fitzroy Crossing (Fitzroy),
4.1 in Nauiyu Nambiyu (Daly), and 5.3 in Kowanyama (Mitchell). The
number of residents per non-indigenous household in the same
locations was 2.1, 2.2 and 1.6 people respectively (ABS, 2006).
2.3 Environmental management context
Grazing is the most extensive land use in all catchments whilst
resurgence in mining activities is underway (Webster et al., 2009).
Water is extracted for agriculture, albeit at relatively low rates,
although there is a potentially large future demand in the Daly and
Mitchell River catchments. Recent plans to dam and divert water
resources or subdivide pastoral leases, clear land and irrigate
farms in the absence of comprehensive environmental assessments or
regional natural resource management plan generated considerable13
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controversy in the Fitzroy and Daly River catchments (see Jackson,
2006; Toussaint, 2008).
Two environmental assessment processes are commonly applied to any
proposal to develop or significantly impact water resources and
both require consideration of the socio-economic effects of
environmental change. First , public works and private project
decisions that are likely to have a significant environmental
impact are assessed at a site specific level under environmental
impact assessment (EIA) legislation. Under Australia’s federal
system, the regulation and development of natural resources are
primarily the responsibility of Australia’s state governments and
so assessment procedures vary in scope and requirements across our
study sites. Social impact assessment (including consideration of
cultural heritage and other indigenous matters) is required by
regulatory approval processes for resource developments in most
Australian jurisdictions (Franks, 2012) and is often conducted
under the auspices of an EIA.
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Second, environmental flow assessments (EFAs) have emerged as key
mitigation measure in response to the need to meet the water
requirements of the environment and sustain values dependent on
healthy aquatic systems (Gardner, 2006) . Environmental flows are
defined as the ‘quantity, timing and quality of water flows required
to sustain freshwater and estuarine ecosystems and the human
livelihoods and well-being that depend on these ecosystems’ (Pahl-
Wostl et al., 2013: 2). EFAs are now mandatory under Australian
national water policy which also expects that the impacts of water
use decisions on indigenous communities will be accounted for in
water plans. To date, few plans have achieved this objective
(National Water Commission, 2011).
Previous work conducted by two of the authors (Finn and Jackson,
2011) summarised existing EFA processes in Australia and for the
following reasons demonstrated the invalidity of the prevailing
assumption that environmental flow requirements can serve as an
adequate surrogate for the protection of indigenous values:
Chosen targets for flow protection do not tend to include key
species used by indigenous people. In the Daly River for15
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example, the relatively rare Pig-nosed Turtle is the species
of most concern to conservation groups, while the highly
abundant Long-necked Turtle is very commonly caught and eaten
by indigenous people. Should Pig-nosed Turtle be chosen for
flow protection, and those flows not adequately protect the
floodplain habitats that Long-necked Turtle require,
indigenous hunters and fishers could be substantially
impacted.
Chosen flow management objectives do not adequately encompass
indigenous needs. Continuing on from the example above, Long-
necked Turtle are abundant and widely distributed. They are of
next to no conservation concern, so a flow alteration that
reduced their distribution somewhat and resulted in only a
moderate population decline may be quite acceptable to many
stakeholders. However, a moderate decline of a very common and
widespread subsistence species can have a social and economic
impact.
EFAs do not adequately address intangible indigenous social
and cultural relationships with water (see Finn and Jackson,
2011 for examples).16
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The initial EFA conducted in the Daly River for example conflated
environmental and indigenous value (Finn and Jackson, 2011). Over
the course of a number of years, water management objectives were
refined to explicitly include strategies for the protection of
environmental and cultural values. In the current draft Oolloo
Aquifer plan, water is valued for its contribution to the ‘condition
of places that provide physical and spiritual fulfilment to
Indigenous people’ (NRETAS, 2012: 11). However, the plan was
written in the ‘absence of known water requirements and specific
site locations which are important for Indigenous cultural purposes
and recreational activities’ (NRETAS 2012: 13).
Determining the environmental flow requirements for indigneous
values in water plans is a national research priority (NWC, 2011)
and is highlighted as a local priority by the water agencies in our
case study regions. For example, the Draft Kimberley Regional Water
Plan (Department of Water, 2010), which includes the Fitzroy River
catchment, acknowledges that ‘water in the Kimberley has important
cultural and social values which will continue to be understood and17
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recognised in planning and management’ (p.12) and identifies a
priority for research on ‘water-dependent values to establish water
regimes for existing ecological, social, cultural and economic
values’ (p. 34). Under this Plan, ecological, social and cultural
values of waterways will be recognised when assessing new
development proposals.
2.3 Method
Our studies sought to understand the spatial and temporal pattern of
resource use; its social, cultural and economic significance to
local indigenous communities and their economies in three
catchments and thereby contribute to improving environmental flow
and impact assessments (Jackson et al., 2011; Scheepers et al.,
2012). The study in the Daly and Fitzroy catchments commenced first
and ran for four years from 2007, whereas the Mitchell River study
commenced in 2009 and ran for three years.
One hundred households were surveyed from five discrete indigenous
communities from the three catchments (Noonkanbah (Fitzroy, WA),
Fitzroy Crossing (Fitzroy, WA), Nauiyu Nambiyu (Daly, NT), Pine18
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Creek (Daly, NT), Kowanyama (Mitchell, Qld), resulting in a total of
1168 semi-structured interviews. Average survey sample sizes
ranged from 16% (17 households) in the Mitchell River, 20% (36
households) in the Fitzroy River and 23% (24 households) in the Daly
River catchment. In the Daly and Fitzroy catchments households were
surveyed twice every three months over a two year period in 2009 and
2010 (i.e. 16 surveys). In the Mitchell catchment surveys were
undertaken once a month over a sixteen month period. The time
between surveys was reduced in the Mitchell to ensure plants and
animals with a short “harvest window” were not missed by a monthly
survey return interval. The number of households surveyed was lower
in the Mitchell catchment due to (i) competing demands on the local
indigenous researcher resident at Kowanyama; (ii) difficulties
experienced in accessing a vehicle for conducting interviews, as
well as (iii) a busy calendar of social and cultural activities.
A range of questionnaire and interview approaches can be used to
collect economic data on the use of wild resources. Intensive
methods such as direct observation (Altman 1987) and doorstep
accounting (Godoy et al. 2000, Sheil and Wunder 2002) can provide19
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accurate and detailed information. However, these methods are
expensive to implement over large areas and the presence of
observers may influence the frequency and location of harvesting
trips. Less intensive methods such as resource user diaries (Gram
2001) or periodic household surveys of resource use (Gray and Altman
2006) can be undertaken inexpensively over large spatial scales but
are subject to shortcomings related to poor participant recall or
inconsistent or unreliable data collection (Gram 2001, Gray et al.
2005). We used household surveys repeated over short time periods
to quantify harvests and catch.
Households participating in the survey were not selected at random.
Due to the frequent timing of repetitive interviews and their
potentially intrusive nature (Altman, 1987), we asked community
leaders to nominate households likely to be interested in
participating. In the manner of the snowball technique, households
initially chosen were asked to suggest others to participate
(Goodman, 1961). To reduce bias, we requested respondents to
nominate households that harvested resources from different areas
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or at different rates of intensity (for further detail see Jackson
et al. 2011; Scheepers 2012).
The economic value of resources consumed by households was
calculated using the replacement cost method (Asafu-Adjaye, 2005)
where the market price of a proxy or substitute is used for products
without a market value:
GRV = (aWt*aTN*aV) + (bWt*bTN*bV) + (cWt*cTN*cV)
Where:
GRV = Gross replacement value;
aWt; bWt; cWt = mean weight in grams of large (a), medium (b) and
small (c) plants or animals;
aTN; bTN; cTN = total number of individuals of a species
harvested or consumed in the large (a), medium (b) and small
(c) size classes;
aV; bV; cV = “shop value” of the replacement item for a species in
the large (a), medium (b) and small (c) size classes.
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While this method provides only a partial economic value and may not
capture other non-material values, it can be readily compared with
existing information on household cash incomes, such as census
data . Data from the 2006 census conducted by the Australian Bureau
of Statistics (ABS, 2006) is used to report results here because it
was used as the basis for designing the research program and
selecting sample sizes. We report our replacement values as a "per
household, per fortnight" figure to enable such comparisons with
cash income and to give the reader a sense of scale when considering
our findings.
Replacement items were chosen to reflect the predominate use of the
species they were to replace (e.g. food or bait), as well as a strong
or weak preference for the species. Barramundi ( Lates calcarifer ), for
example, was considered by respondents from the Mitchell River as a
prized catch, and so its value was equated with a more expensive
store-bought item such as ‘Yellowtail Kingfish’ ( Seriola lalandi),
priced at A$39.45/kg. Likewise, Long-necked Turtle, a commonly
eaten species of turtle, was allocated the price of T-bone steak
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($20.69), while Pig-nosed Turtle (often viewed as ‘special’) was
allocated the more expensive value of fillet steak ($35.50).
Prices for replacement items were obtained from a major supermarket
chain with an online catalogue for goods and prices as product
availability in local stores was usually too low to obtain useful
replacement items for all species. Very high prices are charged for
goods in the local store because of geographic remoteness. The use
of an online catalogue should therefore be viewed as a conservative
approach to the final valuation.
Once replacement values were calculated, data were reviewed to
assess patterns of resource use and the resulting value of
consumption. For each catchment we looked at the 10 species of
highest value that were consumed (per fortnight/household), giving
some insight into how replacement value was spread across species.
We then calculated the top 5 sites in each catchment by replacement
value, and which species within each of these sites contributed to
this value. This information enabled us to assess whether resource
use value was spread across a large or small number of sites, and23
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whether the economic value of each site was well defined by a small or
large suite of harvested species.
3.0 Results
The full inventory of aquatic resources harvested shows that fifty-
six species were harvested from across the sites, with twelve
species being common to all three (Appendix 1). The total value of
consumed wild resources to surveyed communities ranged from
approximately A$36 per household per fortnight in the Fitzroy
catchment to almost three times that amount in the Mitchell River,
where the consumption was valued at $99.99 (Table 1). Long-necked
Turtle ( Chelodina rugosa ) was the highest value species consumed by
survey households in the Daly catchment. Non-fish species made up a
higher proportion of resources commonly consumed in the Daly than
they did in the Fitzroy or Mitchell catchments (Table 1).
Seven of the ten species with the highest replacement value consumed
in the Fitzroy catchment were fish species (Table 1). Two of these
fish species, Bony Bream ( Nematalosa erebi ) and Spangled Perch
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( Leiopotherapon unicolor ), are small bodied species predominately used
as bait for catching other fish. Black Bream ( Hephaestus jenkinsi ) and
Catfish ( Arius spp.) are two popular food species that were of high
value to indigenous households in the Fitzroy. Freshwater Sawfish
( Pristis pristis ) was harvested in lower numbers than other species but
their large size (up to 2m) mean that each individual has a high
replacement value. Cherabin ( Macrobrachium spinipes ), also known as
freshwater prawn, is a large-bodied decapod crustacean that is
captured from rivers and billabongs. Smaller sized Cherabin are
used for bait, while the larger bodied individuals are eaten.
Cherabin are found in much larger sizes later in the dry season.
These results confirm the importance of the river channel to
indigenous resource use in the Fitzroy and highlight the need to
sustain required flows throughout the year.
Insert Table 1 here
In the Mitchell, the popularity of freshwater fish species like
Catfish, Barramundi, Black Bream and Freshwater Jewfish (Neosilurus
spp.) reflects the frequent use of freshwater creeks and waterholes25
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at Kowanyama (Scheepers et al., 2012). However, the relatively high
value of saltwater species such as Grunter ( Pomadasys kraakan ) and
Threadfin Salmon ( Polydactylus macrochir ) in the Mitchell suggests that
the higher level of access to saltwater and estuarine environments
by survey respondents here than in the Fitzroy and Daly can have a
substantial effect on which species are harvested and consumed.
The ten species with the highest combined replacement value
accounted for more than 95% of the total value in all three
catchments (Daly River: 97%; Fitzroy River: 98%; Mitchell River:
95%). While the ten species with the highest value represented more
than 97% of the total value recorded in the Daly surveys, the same was
not true when data was analysed by site. The five sites providing the
highest value of consumed resource in the Daly River (Table 2)
provided only 53% of the total replacement value. Within sites
however, the five species contributing the highest replacement
value at each site represented more than 99% of the total value
derived from that site (Table 2).
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The ability of a subset of sites to represent total replacement
value was even more limited in the Fitzroy River catchment. The five
sites representing the highest replacement value of household
consumption during our surveys represented only 32% of the value for
all sites in the catchment (Table 2). As was true in the Daly River
catchment, the five species contributing the highest replacement
value within each site represented well over 90% of the site’s total
value (Table 2).
Insert Table 2 here.
In the Mitchell, the five sites providing the highest household
consumption value accounted for only 37% of the total value (Table
2). Within sites, the five species contributing the highest values
represented more than 93% of the total value for those sites. This
suggests that people are actively selecting these sites because
they support high-value species, and that the range of species
harvested from each individual site is not large.
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The use of household consumption data in this study allows for an
estimation of the value of customary use of aquatic resources as a
proportion of indigenous household income (Table 3). Household
incomes were derived from 2006 census data (ABS 2006) for the “Urban
Centres and Localities” of Fitzroy Crossing (WA), Nauiyu Nambiyu
(NT) and Kowanyama (Qld). Each of these townships represents the
major urban centres in our study areas. Median fortnightly
indigenous incomes ranged from $1336 (Nauiyu Nambiyu), to $1938
(Kowanyama). Fitzroy Crossing had a median fortnightly household
income of $1640. The value of household consumption of aquatic
species in each location ranged from 2.2% to 5.2% of the median
fortnightly income (Table 3).
Perhaps more tellingly, the imputed value of household consumption
of bush foods (non-market) represented a substantial proportion of
the total value of household food consumption (non-market and
market/store bought). A parallel study to ours collected data on
household expenditure on store bought foods in the Daly River and
Mitchell River catchments (Stoeckl et al., 2011). When the total
value of food consumed by indigenous households in these two28
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catchments was compared with the imputed value of bush foods, the
percentage contribution of bush foods represented 13.0 % of the
total in the Daly, and 22.7% of the total in the Mitchell (Table 3).
Insert Table 3 here
4.0 Discussion
While largely unaccounted for in official estimates of food
production, subsistence activity is a key feature of the remote
indigenous customary economy (Altman, 2001), providing households
with a low-cost means of supplementing low incomes in economically
disadvantaged regions (Biddle, 2009). Financial stress is widely
reported across north Australia with, on average, close to 30% of
indigenous households stating that they had experienced
difficulties paying for everyday necessities in the 12 months prior
to interview, including food, clothing, medical bills and housing
costs ( 32.6% of the NT’s indigenous population, 29.1% of Western
Australia’s and 23.9% Queensland’s) (ABS, 2008). With wild harvest
accounting for between 13% and 23% of the value of food consumed in
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indigenous households in the Daly and Mitchell catchments (and
presumably offsetting the purchase cost of food by the equivalent
amount), it is clear that the long-standing reliability and ready
availability of these species contributes important food supplies
to these communities.
In each of our study areas, the top 10 species made up more than 95% of
the replacement value of the subsistence harvest. A focus on those
species that provide the greatest subsidy to indigenous household
incomes constitutes a useful way of overcoming the conventional
exclusion of indigenous values from water resource decisions and
mitigation actions (Finn and Jackson, 2011). Results from the Daly
case for example influenced a quantitative environmental flow risk
assessment that informed the Oolloo Aquifer Water Allocation
Plan (using Bayesian Belief Network predictive models; Chan et
al., 2012). Researchers learnt that two species at high risk
from low flows, black bream and barramundi, were amongst the
most economically important of species to indigenous
communities, resulting in their inclusion in the risk
assessment. In the absence of this information, the assessment30
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would have been conducted on barramundi alone, a species that
is not as popular with indigenous fishers as it is with
recreational fishers.
Our results have two key implications that provide direction for
environmental managers with scarce resources to devote to assessing
and monitoring environmental change and impacts on indigenous
resource use throughout a catchment. First, they can focus on a
subset of economically valuable species in efforts to fill current
gaps in knowledge of hydrological discharge, ecological response
and subsequent impacts to indigenous harvest and other values. With
a sound understanding of these relationships the level of exposure
of these popular species to adverse risks from water resource
development and the consequences for local livelihoods could be
assessed within a social component of an EIA or environmental flow
assessment. With information on life histories and flow ecology,
individual species of high value could be selected as integrated
indicators potentially capable of revealing the ecological and
socio-economic effects of anticipated changes in hydrological
flows. For example, any flow alterations that impact Long-neck31
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Turtle and reduce indigenous catches are likely to have an adverse
effect on indigenous household incomes and well-being in the Daly
River catchment because Long-neck Turtle alone made up 39% of the
total replacement value (Table 1).
Second, in light of these results, managers can more confidently
target high value species in management and mitigation actions than
if they were to employ a spatial approach. In contrast to the
contribution of the top 10 species to replacement value (> 95%), the
contribution of the 5 highest value sites in each location ranged
from only 32% to 53%. For example, managing flows in the Fitzroy for
the single species with the highest replacement value (Black Bream)
would capture 35% of the total value. Whereas, managing flows for
the single highest value site (Fitzroy River down from Noonkanbah
Airport), would capture only 12% of total replacement value. While
we are not recommending that single-species or single-site flow
management strategies should be considered, our data certainly
suggests that consideration of a targeted subset of species is
likely to capture a larger proportion of the replacement value of
indigenous harvest than targeting a subset of visited sites.32
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A further word of warning on focussing entirely on species and
excluding valuable spatial information is warranted. In the
Mitchell catchment, survey respondents targeted key species at
sites where they knew they could be successful, but also consciously
utilised a high number of peripheral or secondary sites and habitat
types across their customary estates. This strategy was described
as an appropriate cultural protocol; those who did not adhere to
these practices were considered ‘greedy’ for concentrating their
efforts on a select few sites perceived to be of higher risk of over-
exploitation and their behaviour was viewed as ‘disrespectful’ to
the ancestors, the creators and providers of these resources
(Scheepers, 2012). To account for this cultural driver of site use,
our study further endorses a regional approach that could be
implemented in a stratified fashion, with the ‘high value’ sites
subject to more intensive monitoring that other sites.
Notwithstanding the advantages of the approach described here for
prioritising mitigation and management actions for development
proposals, care needs to be taken when assumptions are made about a33
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subset of ‘high value’ species. ‘Value’ is a culturally complex,
context-dependent concept that requires an approach capable of
encompassing multiple valuation criteria (Adamowicz et al., 1998;
Toussaint, 2008; 2010) in addition to respecting the views of
particular cultural groups to substitution, commensurability and
replacement (Chan et al., 2012). For example, for one Daly River
indigenous language group, the death of a man whose totem was the
Magpie Goose resulted in cultural restrictions on the consumption
of that species for a substantial period of mourning. Cultural
practices may well affect the weighting given to goods or resources
that are assumed to be substitutes and, by this logic, which classes
of value are amenable to trade-offs in analytical frameworks such as
benefit-cost analysis. Indeed, the use of economic proxies to
quantify the value of aquatic resources as a function of their
substitutability was foreign to many resource users in our study
sites.
As prices and therefore economic values vary over time according to
demand and supply functions outside the study area, it is plausible
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that were this work repeated in 20 years findings might show (a) the
replacement cost of harvest as a percentage of household income has
changed and/or (b) the replacement cost of particular species has
changed. This does not alter our finding that there is benefit in
targeting a small subset of species that make up a large proportion
of economic value. While economic values and perhaps the subset of
high value species may change over time, the need to understand how
flow alterations may affect key species and subsequently impact
local communities will remain high for as long as indigenous
communities wish to pursue their distinct lifestyles. A baseline
data set is a prerequisite to exploring whether future observed
changes are a result of fluctuating market prices, household income
levels or rates of catch per unit effort resulting from
environmental perturbation. Knowledge that a small group of species
can adequately describe the potential socio-economic impact of
water management decisions may motivate practitioners to adopt a
more socially inclusive approach to assessment and monitoring
programs.
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5 .0 Conclusion
The subsistence use of aquatic resources makes a valuable
contribution to indigenous economies in the three regions under
study. Water resource developments that alter river flow regimes,
modify habitat availability, restrict access and influence species
distributions could reduce fishing and harvesting rates, in turn
affecting indigenous livelihoods and well-being (Stoeckl et al.,
2013). Yet the subsistence value of ecosystem goods and services has
not been recognised through quantification in any Australian
assessments of environmental flows or other water use decisions
(Finn and Jackson, 2011), nor in EIAs relating to development
projects like the McArthur River mine in the Northern Territory
where the river channel was diverted to allow open cut mining.
Rather than emphasize the trade-offs between non-market and market
values, for example wild food versus commercial crops, we reveal the
importance of wild aquatic resources relative to one another by
providing an ordinal scale of value. A very large proportion of the
gross replacement value that we calculated for aquatic species was
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made up of a relatively small subset of high value species. These
species could be targeted for environmental water risk assessments
and as integrated indicators of change in river flows, ensuring that
the monitoring of flow perturbations account for some indigenous
values. However, the total value of species harvested was
distributed across a large number of locations; a feature to note
for assessment of water resource use impacts. While a relatively
small subset of species will encompass a large proportion of the
value of customary harvest, a subset of aquatic sites will not.
Scoping for indigenous values in flow assessments should aim to be
spatially inclusive, but can focus on small subsets of key species.
The valuation undertaken in this study serves three purposes.
Firstly, it provides preference weights for species and sites of
importance for indigenous subsistence, and, secondly, it assists
water managers who have limited resources and knowledge to target
areas of greatest economic value to indigenous communities in their
research, management and mitigation activities. Lastly¸ we find it
a useful tool for establishing a baseline from which to measure
socio-ecological system changes. As argued by Emerton (2005),37
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linking physical changes in ecosystem status and integrity to
changes in economic values and expressing the results as indicators
or measures allows such values to be integrated into broader
economic impact analysis (see also Stoeckl et al., 2013).
Substantial gaps in knowledge remain, ranging from the flow
requirements of species harvested by indigenous people to the
relationship between species’ population dynamics and indigenous
harvest success. These gaps mean that clear calculations of changes
in the replacement value of aquatic species associated with flow
alteration scenarios remain problematic. However, studies like
this lend legitimacy to a targeted approach to filling those gaps
and advancing the inclusion of socio-economic and cultural values
in water resource planning and management and in environmental
impact assessments.
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Acknowledgements
The paper is based on research conducted under the Tropical Rivers
and Coastal Knowledge Research Hub (TRaCK). TRaCK receives major
funding for its research through the Australian Government’s
Commonwealth Environment Research Facilities initiative; the
Australian Government’s Raising National Water Standards Program;
Land and Water Australia; the Fisheries Research and Development
Corporation and the Queensland Government’s Smart State Innovation
Fund.
The authors would also like to acknowledge the contributions of
traditional owners in the Daly River, Fitzroy River and Mitchell
River regions and their representative organisations, the Northern
Land Council, the Kimberley Land Council and the Kowanyama
Aboriginal Land and Natural Resource Management Office.
Constructive input to various aspects of the research was provided
by Jon Altman, Michael Douglas, Brad Pusey, Mark Kennard, Natalie
Stoeckl, Emma Woodward, Alan Andersen, Marcus Barber, Tony
Griffiths, Sandy Toussaint and Geoff Buchanan. We would39
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particularly like to thank Michael Douglas and Vanessa Adams for
their helpful comments on an earlier draft of the manuscript and the
reviewers. Any errors and omissions are the responsibility of the
authors.
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