MANAGING NITRATE LEACHING TO
GROUNDWATER: AN EMERGING ISSUE FOR
Raymond Ford
Ken Taylor
Environment
Abstract
Most of
A comprehensive review of groundwater nitrate-N data, carried out by Environment Canterbury in 2002, found that in areas where rivers and streams were the dominant source of recharge to groundwater and in the coastal confined gravel aquifers, nitrate concentrations were low. In areas where groundwater was shallow, unconfined, and recharged primarily by soil drainage, nitrate concentrations were significantly higher. The nitrate probably originates from a variety of sources, including past and present agricultural land uses, community sewerage schemes, industrial discharges to land, and septic tanks. The predominant trend is for increasing nitrate concentrations, mostly in wells on the lower, or eastern, parts of the plains.
Nitrate concentrations in coastal groundwater-fed streams are high, and exceed the guidelines for the control of periphyton growth, but are still mostly below the guidelines for the protection of aquatic ecosystems.
Environment
The trend of increasing nitrogen loadings
on the unconfined aquifers, spring-fed streams and coastal waters, as a result
of the cumulative effects of land uses is potentially a significant issue for
the region and may require catchment scale limits on nitrogen loads.
Environment
1. Introduction
Over the last fifteen years there has
been a major increase in agricultural production in New Zealand, while at the
same the time there is increasing evidence that the country’s freshwater
resources are becoming nutrient enriched and degraded as a result of pollution
from non-point sources (Parkyn et al., 2002; PCE, 2004). If current
trends persist, further decline in water quality is inevitable. Nationally,
this issue is foremost among public concerns about the environment, and there
is increasing pressure for public agencies to reverse the trend (PCE,
2004). This growing public unease is
also mirrored in
The
There is now a greater recognition of the linkages between groundwater quality and land use, and particularly of the potential impacts of nutrient leaching from farming activities. This recognition has focused attention on the state of the groundwater resource, and the science and policy required for its sustainable management (Lincoln Ventures, 1997).
2. Background
Groundwater
in the
The
aquifers receive a continual recharge of water from the alpine rivers
traversing the plains. This is supplemented by a variable recharge from
rainfall. Water percolates downwards through the gravels and then flows
laterally towards the coast where it emerges as springs and seeps, in the beds
of many small streams on the lower reaches of the plains, in the beds of
shallow coastal lakes such as Lake Ellesmere/Te Waihora (
Figure 1.
Economic value of
groundwater
Groundwater is an important natural and
economic resource for the region. Of all of the water allocated for
out-of-stream uses in
The Christchurch-West Melton groundwater system, the “Ashley Downs” and the Central Plains aquifers, which underlies about half the area of the Canterbury Plains, have been ranked among the top five aquifer systems in New Zealand in terms of their economic value as a source of water for domestic and industrial uses (White et al., 2004).
For many farming operations, irrigation has shifted over the last 15 years from a “drought proofing” operation to become an essential input to satisfy market demands relating to the quality and quantity of agricultural products (MAF, 2004).
Over a similar period, land uses in the
region have continued to develop and intensify, in some instances at an accelerating
rate. Between 1985 and 1999, the area of irrigated land in
Since 1985, there has been
an enormous increase world-wide in the use of inorganic fertilisers. About half
of all the inorganic fertiliser used on the planet has been used during this
period (NRC, 2001), a trend which has also occurred in
Groundwater is the principal
source of drinking water in the region.
3. Groundwater quality
Most of
Table
1: Principal aquifer types in the
|
Aquifer Type |
Water quality
characteristics |
Susceptibility to
contamination |
|
Shallow unconfined or
semi-confined aquifers |
Water
quality variable, influenced by geology and overlying land uses |
High risk, because of: · shallow water table · thin soils and relatively low proportion of
organic matter with low capacity to assimilate contaminants · Gravel and sands overlying water table
are highly permeable and are comprised of inert greywacke gravels that have
low capacity to remove contaminants from water. |
|
Deeper parts of unconfined
or semi-confined aquifers |
Generally
high water quality. Very low concentrations of nutrients and microbiological
contaminants. Water quality in some areas may be affected by local geology,
e.g. buried peat deposits can increase iron and manganese concentrations. |
Moderate risk. Depth to the water table provides some
protection, but deep groundwater may be vulnerable to contamination from
persistent or mobile contaminants, or land use activities in inland recharge
areas. Within semi-confined layers, upwards pressure gradients and lower
permeability confining layer may provide some degree of natural protection. |
|
Coastal confined gravel
aquifers |
Generally
very high water quality. Water quality in some areas may be affected by local
geology, e.g. buried peat deposits can increase iron and manganese
concentrations. |
Relatively low risk. A combination of upwards groundwater
pressures and layers of fine sediments. Over-abstraction could reduce
groundwater pressures and cause downwards movement of contaminants or lateral
salt water intrusion. Land use activities in the recharge area may threaten
groundwater quality. |
|
Non-alluvial
aquifers |
Low yields
of water. Water quality is variable and is strongly influenced by local
geology. |
Relatively low risk, but dependent on the nature of the
fracture system of the parent rock. Contaminant movement is difficult to
predict. |
Groundwater monitoring
Since 1986 Environment Canterbury (and its predecessor, the North Canterbury Catchment Board) has undertaken a regular sampling programme to assess groundwater quality in the region. A wide range of determinands are analysed, including major ions, microbiological indicators, nutrients, metals, trace elements, pH, and conductivity. The results are analysed with reference to the Drinking Water Standards for New Zealand 2000 (MoH, 2000), and published annually by Environment Canterbury (e.g., Abraham & Hanson, 2004a).
A comprehensive review of groundwater nitrate-N data held by Environment Canterbury was carried out in 2002, prompted in part by public concerns about the effects of new irrigation schemes and intensifying agricultural land uses on groundwater quality (Hanson, 2002). More detailed surveys were subsequently undertaken to define areas where nitrate-N concentrations exceeded the Ministry of Health’s Maximum Acceptable Value (MAV) for drinking water (Hayward & Hanson, 2004; Abraham & Hanson, 2004b). This value is a concentration of 50 mg/L nitrate, which is equivalent to a concentration of 11.3 mg/L nitrogen as nitrate-N (MoH, 2000).
Under natural conditions, nitrate-N concentrations in groundwater are low, generally less than 3 mg/L (Hanson, 2002), and possibly as low as 1 mg/L (Close et al., 2001). Generally, low concentrations occur in groundwater where rivers and streams are the dominant source of recharge, and in coastal confined gravel aquifers between the Rakaia and Ashley rivers, where river recharge predominates and the confining layers of fine sediments protect the aquifers from direct contamination from the land surface. In other areas, away from the coastal confined gravel aquifers and areas dominated by river recharge, nitrate-N concentrations are generally higher (greater 3 mg/L) indicating that human activities are affecting groundwater quality (Figure 4).
Nitrate-N
concentrations in shallow, unconfined groundwater fluctuate seasonally,
between 2-6 mg/L, with the higher
concentrations occurring during winter or spring, and the lower concentrations
during autumn (Hanson, 2002).
In the unconfined aquifers north of the
In the Central Plains, between the
Rakaia and Waimakariri rivers, high nitrate-N concentrations occur south and
west of
In the southern part of the Canterbury Plains, Ashwick Flat near Fairlie, and the lowland areas south of Timaru, the average concentrations of nitrate-N range between 2 and 8 mg/L, with areas of locally high nitrate concentrations exceeding 0.75 MAV. Nitrate contamination originates from multiple sources, including agricultural activities in the catchment, land discharges from industry, and small discharges to land, such as those from domestic septic systems ( Hanson 2000, Abraham & Hanson, 2004b).
In the vicinity of Ashburton,
concentrations in shallow groundwater are high for
The source of nitrate contamination in the Ashburton District is attributed to the cumulative effects of agricultural land uses, compounded in some areas by large industrial discharges to land (Abraham, & Hanson, 2004b; Hayward & Hanson, 2004). Land disposal of effluent from meat processing plants north-east of Ashburton has created two long plumes of nitrate contamination in groundwater down-gradient of the disposal sites (Hayward & Hanson, 2004). Nitrate-N concentrations in many of the wells within the plumes are above the limit set by the Drinking Water Standards for New Zealand 2000 (MoH, 2000).
The highest concentrations of nitrate-N occur in wells less than 50 m deep, but
there is no clear relationship between well depth and nitrate concentration
(Figure 5) (Hanson, 2002). In wells deeper than 50 m, nitrate-N concentrations
tend to be lower than for wells less than 50 m, and are generally less than the
MAV. However, concentrations in some deep wells, including wells deeper than
100 metres, are near or exceed 0.5 MAV.
Figure 6. Trends in
nitrate-N concentrations 1995-2004
There appears to be a long-term trend
of increasing nitrate-N concentrations in groundwater (Figure 6). Of the 212
wells analysed by Environment Canterbury for trend, 40 (19%) showed increasing
concentrations while in 19 wells (9%), the trend was for decreases over time
(Environment Canterbury unpublished data[1]).
Most of the wells with increasing trends were generally distributed across the
lower, or eastern, part of the Canterbury Plains. Increases in nitrate–N
concentrations with time have been observed in other places with unconfined
aquifers, such as the lower Waitaki, Ashwick Flat, Culverden and Kaikoura,
although in these areas the dataset covers a shorter period (5 -7 years). Wells
with decreasing nitrate-N concentrations have been identified in a number of
places on the Canterbury Plains; of these, five were located in an area of
Age of the groundwater
Studies of groundwater in the central and northern parts of the Canterbury Plains generally place the age of the groundwater between 30 and 70 years, particularly in areas where rainfall recharge predominates. Younger groundwater is found in those areas where there is recharge from streams and rivers (Stewart et al., 2002).
A considerable amount of time may elapse between a land use activity and the appearance of nitrate-N in a well. The effect of current land uses on groundwater quality may not become apparent for many years. Measurements of the nitrogen isotope N15 suggest that the increase in nitrate-N concentrations in groundwater began around 1950 with the post-war intensification in agriculture (Stewart et al., 2002). Some of the measured concentrations of nitrate-N in the groundwater may have originated from land use activities several decades ago, before large areas of irrigation were developed. If rural land uses continue to intensify and more dry land areas are converted to irrigation and intensive dairying, nitrate concentrations can be expected to increase further (Hayward & Hanson, 2004).
Spring-fed streams of the
lower plains
Spring-fed streams draining the lower or coastal parts of the Canterbury Plains are nutrient enriched with excessive concentrations of nitrate (Meredith & Hayward, 2002).
The median concentrations of nitrate-N in coastal streams draining the plains between the Rangitata and Rakaia rivers are above 5.6 mg/L (0.5 MAV) although mostly below the guidelines for the protection of aquatic ecosystems (Meredith et al., 2005). All of the streams greatly exceed the national guidelines for the prevention of excessive periphyton (algal) growth (Biggs, 2000). Enriched groundwater, as a result of irrigation and land use activities further up the plains, is likely to be the principal source of elevated nitrate-N concentrations (Meredith et al., 2005). Under low or base-flows, the annual loss of soluble nitrogen from these catchments to the coast is estimated to be about 1400 tonnes per annum[2].
Phosphorus concentrations in the spring-fed streams are generally very low, although close to the guideline values for controlling excessive growths of periphyton. The annual loss under low or base-flows is about 4.8 tonnes/ annum or 0.12 kg/ha.
Although the lowland waterways are enriched with nitrogen, excessive plant growths are not widespread, indicating that low phosphorus concentrations are probably limiting plant growth (Meredith & Hayward, 2002; Meredith et al., 2005).
4. Groundwater quality management and the Proposed Natural Resources Regional Plan
Environment
The Proposed Natural Resources
Regional Plan (PNRRP)
The Council decided to implement its
statutory responsibilities by preparing a regional plan to manage the impacts
of human activities on the soil and water resources of the region. Public views
were sought initially on the principal soil and water issues facing the region
(CRC, 1995; ECan, 1999; ECan, 2001a). A draft regional plan setting out
objectives, policies and methods for managing the region’s soil and water
resources was prepared, and released for public comments in October 2001 (ECan,
2001b). Feedback from the community was considered by the Council, and further
technical work was undertaken. The
revised plan (Variation 1) comprising five chapters, was formally notified in
July 2004. The period for public submissions on the plan closed in December
2005, and, as of February 2006, the submissions were being analysed by staff,
with the view to beginning hearings by mid 2006.
Chapter 4 of the PNRRP contains three water quality objectives, which establish specific water quality outcomes for rivers, lakes and groundwater, and sources of community drinking water.
The groundwater quality objectives are based, in part, on the Drinking Water Standards for New Zealand 2000 (MoH, 2000). For aquifers where the groundwater quality is still high, the water quality is to be maintained in that state. Where groundwater quality has been affected by human activities, the maximum concentration of nitrate-N should not increase by more than 2 mg/L above the maximum concentration measured between 1996 and 2001, and should not exceed 11.3 mg/L. In community drinking water supply areas, the maximum concentration of nitrate-N in groundwater should not be greater than 5.6 mg/L.
Each objective is accompanied by policies that set out the position the Council will take on the use, development or protection of a natural resource. The policies are designed to achieve the objective. Underpinning each policy is a suite of regional rules and other methods. In most cases, a combination of methods will be needed, whereby each method complements the strengths and weaknesses of the others.
Broadly, the PNRRP sets out a three-pronged approach to managing nitrate contamination of groundwater, comprising a mix of regulation, education and advocacy, and investigations.
a) Regulation
(i) Point source discharges
The PNRRP proposes that prospective point source discharges to land should be considered from three aspects. Before allowing a discharge to land, the Council must be satisfied that measures have been taken to minimise the volume and concentration of contaminants to be discharged or that the applicant for the resource consent cannot use an existing treatment system or network. If a discharge to land is required, the discharge shall be applied in a way and a rate that matches the assimilative capacity of the soil. If the discharge is likely to result in contaminants reaching groundwater, the contaminant plume must be kept as small as practicable and there must be no adverse effects on other groundwater users.
The chapter also contains a list of community drinking water sources in the region, and includes a method for calculating the size of a wellhead protection zone around a community water supply well. Activities occurring within a well-head protection zone that are likely to pose a risk to drinking water quality will attract a much greater level of scrutiny prior to authorisation.
Nitrate can also enter groundwater as a result of a variety of discharges to land, including those from fertiliser, offal pits, human and animal effluent, and waste from industrial and trade processes. Several of these activities, such as the discharge of animal effluent to land, were controlled by regional rules in the Canterbury Transitional Regional Plan (CRC, 1991). As part of the development of the PNRRP, the existing rules were critically reviewed (LE, 2001; PDP, 2002), and most of the changes to the existing conditions recommended by the reviewers were incorporated into the new regional plan (e.g. discharge of animal effluent[3].
In general, small-scale discharges are permitted under the PNRRP. Authorisation of larger volumes is subject to consideration of the potential adverse effects on a case-by-case basis. However, the use of the aquifers as a direct method of disposing of animal or sewage effluent or hazardous substances is prohibited.
(ii) Managing non-point sources of nitrate
leaching to groundwater
The general approach under the PNRRP is to manage nitrate leaching by using best management practices to minimise losses by matching nitrogen inputs to plant requirements and to prevent the build-up of mineral nitrogen in soils that have a high potential for leaching.
One of the most significant changes proposed is a new regional rule[4] to control nitrate leaching to groundwater in the gravel aquifers from general land uses. A resource consent will not be required (i.e., the discharge will be a permitted activity), but land owners will have to take responsibility to ensure that nitrate leaching from their properties does not exceed pre-determined thresholds.
Figure 7 illustrates the concepts underlying the rule. The two thresholds for nitrate concentration define the points above which a landowner will be required to undertake specific measures to reduce leaching losses. If land use activities result in nitrate concentrations exceeding Threshold 1, landowners must apply best management practices to minimise nitrate leaching. Where land use activities result in high concentrations of nitrate leaching, the landowner will be required to reduce nitrate concentrations to below Threshold 2. In all situations, the average annual nitrate leaching rate for the property will need to be estimated annually to determine the situation relative to the thresholds, and a record of the calculation kept.
Figure 7. Schematic outline
of the nitrate leaching rule
The specific threshold values were determined on the advice of experts in soil-nutrient dynamics, and take into account a number of factors, including drinking water standards for nitrate (e.g. MoH, 2000) and the measured and modelled nitrate leaching losses under different land uses (Bidwell et al., 2003).
Both threshold values represent the average annual concentrations of nitrate-N directly below the plant rooting zone (150 cm). Threshold 1 is set at 8 mg/L, or about 70% of MAV. Land use activities with low leaching rates would fall below the threshold, while annual cropping involving soil cultivation and stocking rates of more than two dairy cows or three beef cattle per hectare are likely to exceed Threshold 1. Threshold 2 is set at 16 mg/L nitrate-N, or 70% of the World Health Organisation limit of 22.6 mg/L above which adverse health effects are observed. High stocking rates, e.g., more than 5 cows per hectare, will most likely exceed this threshold while more typical pastoral land uses are likely to fall below. Arable and vegetable cropping will be distributed above and below the threshold, depending on a range of factors such as crop type, soil drainage characteristics, and farm management systems (Bidwell et al., 2003).
The proposed rule will not take effect until the regional plan becomes operative. It is likely that the rule and the threshold values will be debated and challenged by submitters and different interest groups during hearings on Chapter 4.
b) Education and advocacy
The PNRRP recognises the importance of educating landowners about the
relationship between land use practices and leaching losses, and of practices
which minimise nutrient movement to groundwater. The principal methods in this regard are:
·
promotion of the use of best management practices, whole
farm nutrient management including the use of nutrient budgets, codes of
practice, such as the “Code of Practice
for Fertiliser Use” and the
“Spreadmark Code of Practice”.
- support for further research on nitrogen leaching processes, including nutrient management tools, and the promotion of these amongst land users.
·
encouragement, through education and advocacy, of land
owners, groups or organisations, to modify their land use practices to reduce leaching
losses, in areas where groundwater quality has declined.
c)
Research and investigations - predicting
cumulative effects of land use change on groundwater quality
There are still uncertainties about the effects of nitrate leaching at different spatial scales on groundwater quality from a mix of different land uses, including the time scales over which leaching is likely to impact on ambient groundwater concentrations.
Through the PNRRP, Environment Canterbury has identified a number of topics for future research. The Council is also participating in a project – the “Integrated Research for Aquifer Protection” (IRAP) - with a consortium of regional councils and research agencies. This project will take several years to complete. The outcome will be tools that will allow regional councils to model the cumulative effects of different types of land uses at a range of scales from paddock to catchment, and to predict nitrate concentrations in groundwater at different locations and points in time.
5. Issues associated with
the implementation of the regional plan
The implementation of the PNRRP, and in particular a new regional rule to manage nitrate leaching to groundwater, will be a challenging task. It will require changes within Environment Canterbury and additional resourcing to implement the provisions of the plan, and acceptance from land owners of the need to change long-standing land use practices. The PNRRP represents the first steps at managing nitrate contamination at a regional level. As our knowledge of the problem grows, it is likely other, possibly more far-reaching, measures will need to be applied in the future.
It is likely to be several years, as the plan works its way through the hearings and appeals process, before the nitrate leaching rule becomes operative. During this period, the Council will need to work actively with land owners and industry groups to explain the issue, to illustrate and promote the use of nutrient management tools in general, and to disseminate information on nutrient budgets, in particular.
Undoubtedly there will be resistance to such changes in land use management. There are, however, many advantages to a whole farm nutrient management approach. It offers greater flexibility, compared with a more prescriptive regulatory approach, and allows land owners to optimise production while minimising nitrate leaching by applying best management practices most suited to their individual circumstances. A range of best management practices can be used with the aim of preventing the accumulation of high concentrations of mineral nitrogen in excess of plant needs, particularly before or during winter and early spring when the soil has reached field capacity and drainage rates are high (LE 2001; Di and Cameron, 2001, Ledgard & Menner, 2005).
The application of planning controls can prove difficult when they are at variance with an individual’s economic goals. However, leachate management objectives can be aligned with considerations of profitability. The loss of plant-available nitrogen from a catchment and individual properties can have a significant economic cost (Meredith et al., 2005). The use of nutrient budgeting is likely to lead to significant savings in fertiliser costs. Measures that have both economic and environmental benefits are more likely to be accepted than those that provide environmental gains alone.
In most cases, computer models, which incorporate nitrate leaching, farm management practices and inputs (e.g. fertiliser and irrigation), plant nitrogen requirements, and local environmental characteristics, such as soil type and climate, will be required to provide reliable estimates of nitrate leaching losses. Decision support system models will allow farmers to look at different management options to improve production while reducing nitrate leaching (Di and Cameron, 2002). Estimates of the average annual nitrate loss will depend, not just on accurate input information for the leaching models, but are also likely to require the contribution of people with specialised skills in agricultural nutrient management, to run the models and design appropriate management strategies. While some land managers will have these skills, others may need to rely on outside expertise, such as farm management advisors and fertiliser company representatives.
Models of nitrate leaching fall into one of two categories (Bidwell et al., 2003): nutrient budget models (e.g. ‘Overseer’) which calculate nitrate leaching loss as the residual loss after taking into account all the other inputs or outputs, and nutrient discharge models (e.g., ‘Nitrogen Leaching Estimator” (NLE)) which are empirical models that have been calibrated with experimental data and calculate leachate rates directly.
The use of the models will raise awareness of landowners about the different factors that influence nitrate leaching, while determining whether a land use is above or below the thresholds set by the rule. Potentially, there are likely to be issues of consistency as Council determines whether a particular land use complies with the rule. A particular model may give quite different answers depending on what input values are chosen, or different models are likely to produce results that may be variance with each other.
There has been criticism that existing nutrient models are not suitable for arable or mixed cropping farming. To address this problem, Environment Canterbury is contributing to a programme led by Crop and Food Research Limited with the support of the Sustainable Farming Fund, to develop a nutrient management tool for arable or mixed cropping farmers.
If the nitrate leaching rule becomes operative in its current form, Environment Canterbury will probably focus its efforts, at least initially, on implementing the rule in those parts of the region where groundwater nitrate concentrations are of particular concern. No areas have been identified yet, but the priority is likely to be areas, such as the Ashburton and Selwyn districts where there has been a long history of elevated nitrate concentrations in groundwater, and areas where nitrate concentrations are exhibiting an increasing trend as a result of land use intensification, e.g. Ashwick Flat, near Fairlie.
6. Emerging issues
The coastal environment
In the marine environment, unlike the freshwater environment, nitrogen is the critical limiting nutrient for phytoplankton growth (NRC, 2001). Nitrate is constantly being transported through the region (Figure 8). A substantial quantity of nitrate, generated on the land, eventually finds its way to coastal waters, via spring-fed streams and large rivers, and groundwater emerging at the coast. Coastal waters, because of their location, are the sink for all the nutrients being transported from the land. Estuaries and coastal lakes and lagoons may act as temporary stores or traps and delay the entry of nitrate into the coastal waters.
What are the implications of
large nitrogen inputs for the coastal environment?
Nutrient enrichment of
coastal waters, or “eutrophication”, stimulates the growth of phytoplankton and
can lead to algal blooms, including “red” and “brown tides”, degradation of
habitat, oxygen depletion and the accumulation of toxins in shellfish used as a
human food source. Nutrient enrichment of coastal waters is a significant
problem in the
Algal blooms in coastal waters may be
associated with substantial economic losses. Prohibitions on the harvesting of
farmed shellfish in response to toxic blooms have impacted on aquaculture in
The discharge of effluent or wastewater directly to the ocean also contributes to coastal nutrient loading. In recent years, there has been a shift away from land disposal of wastes to discharges into coastal waters. Christchurch City Council, and Waimakariri and Timaru District Councils are proposing to discharge sewage via ocean outfalls. Several large industries also discharge, or are proposing to discharge, their waste via ocean outfalls.
Further work is needed to understand
the relationship between nutrient inputs to coastal
Table 2 shows the area-specific load to the
Canterbury Bight from the central part of the
Table 2. Total nitrogen loadings to the
|
|
|
|
Baltic Sea 2000 (source EEA, 2005) |
North Sea 2000 (source EEA, 2005) |
|
Catchment
area (million
km2) |
0.026 |
0.167 |
1.6 |
0.53 |
|
Annual
catchment yield (kg N/ha) |
2.35 |
7.92 |
5.4 |
14.4
|
Managing cumulative effects
of land use change on groundwater quality
In
The expansion of intensive agricultural land uses over the unconfined gravel aquifers, such as the conversion from dryland farming to intensive dairying could potentially lead to a serious decline in the quality of the region’s groundwater resources. The trend is likely to be exacerbated if the soils become saturated with nitrogen, resulting in increased leaching of nitrate (PCE, 2004).
An extrapolation of current trends in nitrate concentrations suggests that within 30 years the nitrate concentrations of many shallow unconfined aquifers could exceed the limit set in the Drinking Water Standards for New Zealand 2000 (Hanson, 2002). This conclusion is also supported by modelling of land uses changes at a catchment scale (Di and Cameron, 2002, Di and Cameron, 2004). Already many spring fed streams in the lower Canterbury Plains show elevated nitrate concentrations, an indication of the effects of intensive land use activities occurring further inland. At times of low flows, the median nitrate concentrations in many coastal groundwater-fed streams between Ashburton and the Rangitata rivers are close to the guideline for aquatic ecosystems (Meredith et al., 2005).
Experience from overseas, where diffuse run-off from land uses and point source discharges have contaminated groundwater, has shown that aquifer remediation to drinking water standards is extremely difficult (Freeze and Cherry, 1989). Remediation of aquifers, even if it is technically feasible, is likely to take considerable time and resources, and therefore preventing a decline in groundwater quality is the only effective and economic way of managing the groundwater resource.
One of the major difficulties with managing groundwater contamination is that for most people, aquifers are literally “out of sight and out of mind”: a decline in groundwater quality does not attract the same level of attention until people are personally affected, unlike a river or lake where changes in water quality are readily observed. A feature of nitrate contamination is the time lag, possibly of several decades, between the leaching of the nitrate in the soil profile and its appearance in groundwater (Haynes, 1997). Therefore, the consequences of present land use activities may not become apparent for many years, and once elevated concentrations of nitrate occur in groundwater it may take many years before changes in land use practices bring about any measurable improvement in groundwater quality. This creates a problem when assessing an application for a resource consent and linking the relative contribution of nitrate from the proposed activity to the cumulative effects of land uses on water quality before the activity has commenced and its adverse effects are observed.
At this stage, Environment Canterbury
has not determined how it will manage the total nitrate loads from individual
catchments in the region. The
The allocation of water for out-of-stream uses is an example of one policy approach that could be applied to manage the total nitrate loads. A maximum load or cap for nitrate discharges for each catchment would be determined with respect to some critical value, such as the MAV for drinking water, the toxicity limit for aquatic ecosystems (Hickey, 2002), or the assimilative capacity of coastal waters (NRC, 2001). The total load would then be apportioned among landowners and resource consent holders within the catchment, and could then be traded or assigned to other landowners within the catchment. The allocation of individual entitlements to nitrate discharge, however, raises challenging issues, such as that of determining an equitable basis for calculating the initial allocation to individuals, and, for Environment Canterbury, the practical problems associated with the monitoring of nitrate losses from land use activities and discharges, and administering any system of tradable permits (EW, 2005). Public involvement, including a wide range of groups such as district councils, landowners, research institutes, organisational groups, and Ngai Tahu, in the development of any caps on nitrate loads will be essential if management measures are to be successful.
7. Conclusions
Water quality monitoring has shown that
groundwater in
More fundamentally, we need to improve our understanding of the nutrient fluxes and to acknowledge the linkages between groundwater, spring-fed streams and near-shore coastal waters. Future nitrate policy development will be required to encompass a more integrated or “whole hydrological systems” approach. High groundwater nitrate concentrations are not just a problem for groundwater users. Future management will need to consider the linkages between the different types of water bodies, and to identify those water bodies or values that are the most vulnerable to nitrate inputs. It may be that nitrate management will require the establishment and implementation of maximum loadings for particular catchments, and the allocation of discharge permits in a way that is analogous to the current allocation of water for abstractive use.
What may be particularly challenging to our thinking in this regard, is the prospect that such allocation limits may be determined not on the basis of the potential impacts of land-based activities on other groundwater users or the values in coastal spring-fed streams, but on the implications of those activities for the uses and values of waters overlying the continental shelf, and beyond.
Acknowledgements
Our paper draws on many years of work by the Environmental Canterbury groundwater quality monitoring team. We would like to acknowledge their contributions to our understanding of the issue, and to thank our colleagues for their comments and assistance with the preparation of this paper. Thanks also to Trevor Webb of Landcare Research for his comments.
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