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Assessment of the geochemical environment in the lower Burdekin aquifer: Implications for the removal of nitrate through denitrification Thabonithy Thayalakumaran1, Philip Charlesworth1,2 and Keith Bristow1,2 1CSIRO Land and Water and 2CRC for Irrigation Futures Davies Laboratory, Townsville, QLD 4814, Australia

CSIRO Land and Water Technical Report No. 32/04

August 2004

Cover Photograph: From CSIRO Land and Water Image Gallery: www.clw.csiro.au/ImageGallery/ Description: Sources of electron donors for nitrate reduction in the lower Burdekin aquifer: DOC, ferrous iron and nitrate. Photographer: Renate van Bemmelen © 2004 CSIRO

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Assessment of the geochemical environment in the lower Burdekin aquifer: Implications for the removal of nitrate through denitrification Thabonithy Thayalakumaran1, Philip Charlesworth1,2 and Keith Bristow1,2

1CSIRO Land and Water and 2CRC for Irrigation Futures, Davies Laboratory, Townsville, QLD 4814, Australia

Technical Report No. 32/04 August 2004

Assessment of the geochemical environment in the lower Burdekin aquifer ii

Acknowledgements This work was supported in part by

• CSIRO Land and Water

• The CSIRO Water for a Healthy Country flagship program

• Cooperative Research Centre For Irrigation Futures

We thank the Queensland Department of Natural Resources Mines & Energy (NRM&E) for allowing access to the monitoring bores and providing the pump used to obtain the samples for this study. We also thank Ray Mcgowan, Bob Mooney and Steve Porter of NRM&E for their assistance in sampling the monitoring bores.

We thank a number of Burdekin cane growers for allowing us to sample their production bores.

We also thank Renate van Bemmelen, a work experience student from Hogeschool Zeeland in Vlissingen, The Netherlands, and Jasmine Jaffres, JCU, Townsville, Australia who contributed to the sampling, literature review, and part of the data analysis.

WATER FOR A HEALTHY COUNTRYNational Research Flagship

WATER FOR A HEALTHY COUNTRYNational Research Flagship

Assessment of the geochemical environment in the lower Burdekin aquifer iii

Executive Summary Factors with the potential to affect denitrification in the lower Burdekin aquifer were studied during the September 2003 to January 2004 period. The first sampling occurred during September to October 2003 and the second in January 2004. A total of 57 Queensland Department of Natural Resources and Mines groundwater monitoring bores and a few farmer production bores were targeted for the study. The key findings emerging from this study are:

• The depth to water table (below ground level) ranged from 0.96-10.5m. There was a tendency for more shallow water tables to occur near the coast and deeper water tables to occur inland during both sampling periods.

• Nitrate concentrations ranged from <0.1 to 15 mg/L NO3--N, three times the ANZECC

environmental standard of 5 mg/L. Elevated nitrate levels were mostly found in the Airville-Home Hill areas. Nitrate was undetectable in the nested bores sampled along the coast. This trend was consistent in both sampling periods. Nitrate distribution with depth revealed variable trends between sampling periods. Concentrations were significantly higher in shallow bores during the January 2004 sampling, compared with the September/October 2003 sampling which suggests a quick response to fertilisation and irrigation.

• Ferrous iron (Fe2+) concentration displayed a wide range, from 1 mg/L to as high as 360 mg/L. No well defined geographical distribution was apparent, although some correlation of high ferrous concentrations with geographic area was evident in the Kalamia Mill and Ayr areas. Ferrous levels were below 2 mg/L in the Airville-Home Hill area. Wherever elevated ferrous levels occurred they were mostly found in shallow bores (< 20 M).

• The DOC levels ranged from 4 to 82 mg/L which is very high compared to values reported for other aquifers. High levels of DOC found in groundwater could be a result of carbon from sugars lost during harvest. No specific spatial or vertical distribution pattern was evident, with elevated concentrations of DOC occurring at depths as great as 80 m.

• The aquifer was generally sub-oxic to anoxic, with dissolved oxygen concentration usually < 2 mg/L. Eh levels varied from -120 to +235 mV and pH ranged from 6.0 to 7.6. Observable differences in Eh and DO with depth were not well defined, especially in the nested bores. This could be due to the unconfined and permeable nature of the aquifer.

• A negative correlation was observed between NO3--N and ferrous iron in both sampling

periods suggesting the possible occurrence of denitrification. It is not clear at this stage whether there were only small amounts of nitrate leached into the groundwater or if the nitrate had been denitrified in the groundwater. There was no clear correlation between the DOC and nitrate measured in groundwater. However, the large concentration of DOC combined with the reducing environment in the aquifer indicates a strong potential for nitrate removal through denitrification.

• The negative correlation between nitrate and ammonium also suggests the possible occurrence of dissimilatory nitrate reduction or the accumulation of ammonium as the end product of organic mineralization under aerobic conditions.

• The large apparent potential for nitrate removal through denitrification in the lower Burdekin aquifer may be partially affected by the export of N as ammonium or DON. While beneficial from the point of view of promoting denitrification, the high DOC loading of the aquifer may compromise its role as a water supply by increasing the weathering rates of minerals, especially Fe and Mn hydroxides, leading to higher levels of these metals in the groundwater.

Assessment of the geochemical environment in the lower Burdekin aquifer iv

Future work addressing the potential of this aquifer to consume nitrogen through denitrification should involve:

1. Groundwater sampling to assess the seasonal changes in nitrate

2. Development of greater understanding of the geochemical conditions and the electron donor availability along the groundwater flow path at different depths. This could be used to demarcate nitrate sensitive zones and depths within the aquifer

3. Measuring the rate of denitrification and determining the rate limiting factors (DOC, Nitrate, Ammonium, DO etc)

4. Developing improved understanding of the geochemistry and biology of the aquifer to allow quantification of denitrification in the system

5. Determining the concentration of dissolved organic nitrogen (DON) in the groundwater

6. Determining the lability of DOC in groundwater and the long-term environmental risks associated with a large loading of DOC in the aquifer.

Assessment of the geochemical environment in the lower Burdekin aquifer v

Acknowledgements ii Executive Summary iii 1 Introduction 1 2 Literature Review 3 3 Methods 7

3.1 Site description and aquifer characteristics 7 3.2 Sample collection and analysis 8

4 Results 10 4.1 General patterns in groundwater quality 10 4.2 Geochemical conditions 19

5 Discussion 22 5.1 Major ion chemistry 22 5.2 Redox environment of the aquifer 23 5.3 Nitrate reducing processes 23

6 Conclusion 29 7 Recommendations 30 8 References 31

9 Appendix 1. Data Summary 33 9.1 Appendix 1A. Data summary for the samples taken during September/October 2003 33 9.2 Appendix 1B. Data summary for the samples taken during January 2004 35 9.3 Appendix 1C. Cation and anion data for the samples taken during January 2004 37

Assessment of the geochemical environment in the lower Burdekin aquifer

1

1 Introduction Nitrate is a widespread contaminant in groundwater in agricultural areas, partly because of

intense land-use activities and large scale fertiliser use during the last few decades (Bohlke

et al. 2002). Water containing elevated concentrations of nitrate is unfit for human

consumption (Spalding and Parrott, 1994) and, if discharging to freshwater or marine

habitats, can contribute to algal blooms and eutrophication. The maximum permissible limit in

drinking water is 10 mg/L of nitrate-N (WHO and US Environmental Protection Agency).

Consequently, there has been increasing interest in exploiting any natural processes that

utilise the groundwater nitrate and conditions that enhance the denitrification processes.

While denitrification has been intensively studied with respect to surface water/sediment

interfaces, soil environments and water treatment processes (Barton et al. 1999) it has in

recent years been recognised for its ability to eliminate or reduce nitrate concentrations in

groundwater (Korom 1992; McLarin et al. 1999). Denitrification is a natural process where

nitrate is converted to N2 gas. The process requires a sub-oxic to anoxic environment,

denitrifying bacteria, and sufficient electron donors such as dissolved and particulate organic

carbon (DOC & POC), ferrous iron, and sulphides.

In the lower Burdekin area of northern Queensland the suitability of the deltaic soils for

cultivation, the tropical climate, and groundwater availability has allowed the sugarcane

industry to expand over the last one hundred years, with more than 80,000 ha now under

irrigated sugarcane and other crops (Bristow et al. 2002). Furthermore, this area overlies

major groundwater supplies and because it is close to environmentally sensitive wetlands,

waterways and the Great Barrier Reef Lagoon, its impacts on water quality is under

increasing scrutiny. Nitrate concentrations above the ANZECC guideline for long-term

environmental sustainability (5 mg NO3-N/L) have been reported in 49 out of the 397 sample

bores in the Burdekin region (Weir 1999). During 2000-2002, sampling of ten farmer

production bores in the Ayr or Homehill areas found nitrate levels from 1 to 12 mg/L NO3--N

(Charlesworth et al. 2003). Monitoring of groundwater nitrate concentrations over 2-6 years

however, has shown there has been little change in 90% of wells monitored in the Burdekin

region (Thorburn et al. 2003). In other aquifers groundwater studies have established that

denitrification can be an important mechanism for maintaining or decreasing nitrate levels in

aquifers that are low in oxygen (Postma et al. 1991; Korom 1992; Spalding and Parrot 1994;

Kelly 1999; McLarin et al. 1999). However, the hydro-geochemical environment is not always

suitable for denitrification to occur or the high concentrations of nitrate introduced into

groundwater may exceed the reduction capacity in the aquifer (Postma et al. 1991).

Therefore, for substantial nitrate reduction to occur in the groundwater, favourable hydro-

geochemical conditions must exist or the sediments must have extra reduction capacity.


2

There is ample observational evidence of iron in groundwater in certain areas of the lower

Burdekin including yellowish-red deposits around pump outlets, and pipes blocking due to

deposits. Yellow deposits in pipes or yellowish-red coloured sediment are the indications of

the presence of iron oxide coatings (Puckett and Cowdery 2002). Furthermore, Kelly (1969)

reported the presence of un-complexed ferrous iron, with concentrations ranging from non-

detectable levels up to 100 g m-3 in the Burdekin groundwater. While there have been studies

that have examined the distribution of nitrate or iron in groundwater in separate studies, none

have attempted to link iron and/or DOC and hydro-geochemical conditions with the nitrate in

the groundwater of the lower Burdekin. Investigation of the availability of electron donors

such as ferrous iron, DOC and other supporting hydro-geochemical conditions that enhance

the potential of the Burdekin aquifer for denitrification was the major focus of this study. This

report includes findings from sampling carried out from September to October 2003 and in

January, 2004.


3

2 Literature Review Denitrification refers to a microbial respiratory process where nitrate is used as the terminal

electron acceptor and is reduced to N2 gas (Puckett and Cowdery 1999):

OHNeHNO 223 610122 +→++ +− (1)

Bacteria in aquifers, depending on the species, obtain energy from the oxidation of organic or

inorganic compounds (electron donor). In order to complete this oxidation reaction a

reduction reaction, an electron acceptor (O2, NO3-, Mn4+, Fe3+ and SO4

-) is also required

(Korom 1992). Depending on the availability, bacteria use the electron acceptors in the

following order: O2, NO3-, Mn4+, Fe3+, SO4

- and CH4. This means that when O2 becomes

limited in the saturated zone, bacteria start using nitrate as an electron acceptor. Organic

carbon is the most common electron donor and tends to be oxidised preferentially by

acceptors that yields the most energy to bacteria (Starr and Gillham 1993). The

stoichiometric equation for denitrification by organic carbon follows,

23223 42245 COHCONOHNOC ++→++ −− (2)

However, low DOC concentrations in groundwater and low organic matter content in aquifer

materials tend to limit the potential for organic carbon as electron donors in many aquifers.

Alternatively, other reduced inorganic substrates that may serve as electron donors for

nitrate reduction reaction, including ferrous iron and reduced sulfur from the iron sulphide

minerals such pyrite (Kolle et al. 1985; Postma et al. 1991).

In the case of reduction by Fe2+, the reaction follows,

+−+ ++→++ HFeOOHNOHNOFe 181014210 2232 (3)

The N2 gas produced from the denitrification processes remains in solution until the

groundwater discharges to a surface-water and equilibrates with the atmosphere. Thus an

assessment of the geochemical conditions gives an indication of whether an aquifer has

potential for nitrate removal through denitrification, and also to some level, the fate of nitrate

in groundwater can be predicted (Böhlke and Denver 1995).

Investigations on the heterogeneities in groundwater geochemistry in a sand aquifer beneath

an irrigated field in Illinois, USA showed that denitrification reactions were responsible for

removing nitrate from solution beneath the plume, probably mainly coupled to oxidation of

sulphide minerals. Kelly (1999) reported that the presence of Fe or nitrate was usually

marked by the absence of the other in a shallow unconfined sand aquifer in Mason County,

Illinois. An investigation by Lamontagne et al. (2003) in an aquifer in Wollombi Brook, NSW


4

reported that high nitrate concentrations were only found when ferrous iron was low or below

detection limit. An inverse correlation was found between iron and nitrate concentrations in

groundwater in the unconfined aquifer in Manakau, New Zealand (McLarin et al. 1999).

The choice of the electron donor by bacteria will depend on the supply of these electron

donors and electron acceptors which will be controlled by the aquifer properties, including

dissolved oxygen (DO), redox potential (Eh) and pH. The DO concentrations in groundwater

usually vary with depth. Just below the land surface oxygen concentration is high because of

exchange with the air, and as the depth increases the amount of DO decreases. DO is

thermodynamically preferred over nitrate as the electron acceptor therefore, denitrification

starts only when all the available oxygen has been consumed. Several values have been

reported for the upper DO limit for denitrification to proceed. However, they all mean that low

O2 in the water encourages this process. DO levels of less than 1.7 mg/L and 2.0 mg/L were

reported to be favourable for denitrification by the Minnesota Pollution Control Agency (1999)

and Korom (1992) respectively. Once most of the nitrate is utilised, bacteria start consuming

manganese oxides, iron oxides, sulfate and finally methane as electron acceptors.

The redox potential (Eh) is a measurement of the relative difference in energy between the

oxidants and the reductants present in the environment. In the field it is difficult to obtain Eh

measurements that represent the true Eh conditions due to the presence of multiple redox

couples, which may also not be at steady-state relative to one another. However, Eh can be

used to broadly categorise the geochemical environment in an aquifer. The following

reactions using organic carbon as the electron donor have been listed to occur in the

saturated zone in order of decreasing Eh conditions (Vance 2002),

• Aerobic respiration at +250 mV and higher

• Use nitrate from 250 mV to 110 mV

• Use Mn4+ or Fe3+ from 100 mV to 0 mV

• Use sulfate from 0 to -200 mV

• Use methane from -200 mV and lower

For denitrification using ferrous iron as the electron donor the approximate threshold level of

the redox-potential is +250 mV (McLarin et al. 1999; Spalding and Parrot 1994; Vance 2002).

The Minnesota Pollution Control Agency (1999) defined the sensitivity of an aquifer for nitrate

in three zones; sensitive zone, transition zone and a not sensitive zone (Table 1).


5

Table 1: Nitrate sensitivity zones (Minnesota Pollution Control Agency 1999)

Zone Oxygen (ppm)

Eh (mV) Iron (ppm)

Nitrate (ppm)

Sensitive 6.3 308 0.03 5.60 Transition 4.6 300 0.07 0.50 Not sensitive 1.7 287 0.30 0.05

The chemical or biological reduction of nitrate also depends on the amount of nitrate, the

electron donors (eg.ferrous or DOC) that are present in the groundwater, and the microbial

population. Relative to soils, the time scales of nitrate reduction in groundwater are slow,

mainly relating to the supply of reactants (Postma et al. 1991) and also because most of

the microbes are attached to the porous matrix of the aquifer rather than living freely in

groundwater (Korom 1992). The amount of nitrate and DOC that reaches groundwater

depends on factors such as soil type, soil/aquifer hydraulic characteristics, soil

heterogeneities, seasonal changes, depth of the unsaturated zone and the land use

practices (Böhlke et al. 2002), while ferrous iron levels in groundwater depends on the

aquifer chemistry and aquifer properties. Relatively high concentrations of nitrate and DOC

concentrations have been observed in shallow ground water and they are typically not found

at depth in aquifers (Kelly 1999). While average DOC levels in aquifers are usually below

those necessary to account for the observed reduction in nitrate, DOC levels up to 32 mg/L

have been reported in shallow ground waters of the coastal plain in North Carolina (Spruill,

1997). DOC-enhanced denitirification has been observed to occur in very shallow water table

aquifers and not in aquifers with water tables that are deeper than about 2-3 meters (Starr et

al. 1993). This is due to the longer residence time that the DOC spends in the unsaturated

zone where it is degraded or consumed before reaching the groundwater. Ferrous iron levels

up to 16.7 mg/L have been reported in an unconfined sand aquifer (Postma et al. 1991). It is

clear from the above that an understanding of the hydrology and geochemistry of the aquifer

together with careful sampling time and procedures are required to understand the fate of

nitrate in groundwater.


6

The objective of this study was to characterise the geochemical environment and the

potential for denitrification in the lower Burdekin aquifer. This was achieved by:

1. Analysing the spatial and vertical distribution of nitrate, ferrous iron and DOC in

groundwater.

2. Examining the presence of hydro-geochemical conditions that favour denitrification.

3. Evaluating the potential for denitrification in the presence of ferrous iron and DOC.

4. Using the availability of ferrous iron, nitrate levels and the hydro-geochemical

conditions to assess potential areas for denitrification.


7

3 Methods 3.1 Site description and aquifer characteristics The study was carried out in the lower Burdekin, located approximately 90 km southeast of

Townsville in Queensland. The Burdekin River Delta aquifer comprises sedimentary

deposits in excess of 100 m in some places near the coast which overly a predominantly

granitic basement. The nature of sedimentation is complex with sediments comprising a

mixture of interbedded gravel, sand, silt, mud and clay. The groundwater system in the

Lower Burdekin is generally considered to be unconfined, due to the presence of sandy to

loamy soils and the discontinuous nature of underlying clay layers. The thickness of the

aquiferous sand generally varies between one-third and half of that of the total alluvium

which itself varies from 10-80m (Brodie et al. 1984).

Groundwater flows generally from the south to the north based on potentiometric surface

maps. The hydraulic conductivity varies across the delta from 10 to >300 m/d. From an

assessment of different processes in the groundwater system including infiltration of rainfall,

artificial recharge through pits and channels, river recharge, flooding and irrigation return

flows, it has been simulated that the average recharge rate in the region between 1981- 1997

varied from 330,000 and 650,000 ML/y (Arunakumaren et al. 2000). Water loss from the

aquifer is through discharge to the sea and river, pumping for irrigation and evaporation.

Fertiliser nitrogen recommendation for sugarcane in the Burdekin is 160-220 kg of N/ha/yr

and is usually applied as urea in a single operation at the start of the season (April-October).

Figure 1. Geographic setting of the Lower Burdekin (NBWB-North Burdekin Water

Board, SBWB- South Burdekin Water Board)


8

3.2 Sample collection and analysis

500000 510000 520000 530000 540000 550000 560000Easting

7800000

7805000

7810000

7815000

7820000

7825000

7830000

7835000

7840000

7845000

7850000

Nor

thin

g

Bores <15 m sampled during 2004

Bores >15 m sampled during 2004

Nested bores sampled during January 2004

Bores sampled during September/October 2003

1191020311910204

11910984

11910023

12100012

12100166

1190013111911056

1191027011910038

11910256

11910259

11910260

11910046

12000090

11910196

11910191

11910268

12000204

11910257

12000112

120001141200007912000126

11910808

11910095

1191007311910049

11910124

119102631191004811910036

11910878

11910190

11910975

1191081011910082

11910117

1191074411910066

11910119

11910942

11900045

1190014911910886

11910051

1191024911900150

11910162

11910877

11910150

Figure 2. Sampling locations during September/ October 2003 and in January 2004 Groundwater sampling of monitoring bores was performed in two different periods, one

during September/October 2003 and the other during January 2004. Different bores were

sampled during the two periods. In the first sampling period, 28 single and 2 nested bores

from the delta and BRIA area were sampled along with nine farmer production bores. Twenty

two single bores and 5 nested bores were sampled during January 2004. The monitoring

bores used in this study were installed by the Queensland Department of Natural Resources

Mines and Energy (NRM&E).These monitoring bores were located in the Giru-Baratta, Ayr,

Homehill and Green Swamp areas. The farmer production bores sampled were mostly from

near Ayr. The numbers in the bore names refer to the location of the bore in Figure 2. For the

first sampling we concentrated on bores that had high nitrate or high total iron concentrations

during 1996-2000, as indicated by the NRM & E groundwater database. Bores from inland

and nested bores along the coast were targeted for sampling during the second period. Most

of the chosen bores were confined to sugarcane planted area. Bore numbers 11910263,


9

11910268, 11910257, 12100166 and 12000204 are nested bores each consisting of 4-6

bores installed to different depths. Bores 11910263 and 1210016 were sampled in both

sampling times. The depth of the bores varied from 5.4 m to 88.5 m.

For each bore, the water table depth, screen depth and the type of surrounding vegetation

were recorded. Water samples were withdrawn from monitoring bores using a stainless steel

submersible pump with a Teflon discharge line while the farmer production bores were

sampled directly from the pump outlets. Bores were purged to remove standing water

through a flow cell containing pH electrode. Purging of a bore continued until either three

bore volumes had been pumped out or until the pH had stabilised. Redox potential (Eh), pH

and electrical conductivity (EC) were measured using electrodes in the flow cell to minimise

contact with air. DO measurement was made by continually filling a bottle holding a DO

probe. The pH and Eh measurements were carried out using an electrode equipped with a

Radiometer-Ion check 10. The EC and DO were measured using electrodes connected to

display. All these electrodes were calibrated in the field using standard solutions provided by

the companies. Once these measurements were completed water samples were collected.

Water was collected by inserting one end of the silicon tube connected to the Teflon

discharge line into the bottom of a Erlenmeyer flask in which water was allowed to overflow.

Four sub-samples were collected from each bore for ferrous iron, total iron, nitrate-nitrogen

and for other cation and anions after filtering with a 0.45 µm filter. For ferrous iron analysis,

water samples were collected in vials containing 4.0 ml of 1 M sodium acetate and 2.5 ml of

0.1% phenanthroline (1,10-phenanthroline monohydrate) that complexes the unstable

ferrous iron. For total iron and manganese water was collected in pre-acidified vials. Sub

samples were also collected and filtered for other anions and cations analysis. Eh and the

DOC were not measured in the September/October sampling period. Samples were

preserved in ice packs and transported to the lab.

Chemical analyses were performed at the Davies laboratory or at the Australian Centre for

Tropical Freshwater Research (ACTFR), Townsville. Ferrous iron concentration was

measured by spectrophotometer and the total iron was measured using atomic absorption

spectrophotometry. Nitrate, nitrite, ammonium, sulfate, phosphate and chloride were

determined by standard methods outlined in Rayment and Higginson (1992). Nitrate and

nitrite by cadmium reduction methods, phosphate as soluble orthophosphate by manual

colour method, sulfate by turbidimetric method, chloride by continuous flow analyser system

with a colorimeter, bicarbonate using potentiometric titration. Total dissolved iron,

manganese were analysed using atomic absorption spectrophotometry. DOC was analysed

at the ACTFR using the Shimadzu Automatic carbon analyser. Ionic balance errors for the

analysis were mostly within ±5%.


10

4 Results 4.1 General patterns in groundwater quality 4.1.1.1 Water table The depth to water table, below ground level (bgl), ranged from 2-8.8 m during

September/October 2003 and 0.96 to 10.5 m during January 2004 (Figure 3). The water

table was deep to the south and west of Ayr, but mostly around 7 m. In the Giru – Barratta

region, the water table was mainly between 5 and 6 m depth. The greatest variation in the

water table was noted in the Airville - Home Hill area where both the shallowest (0.96 m) and

the second deepest (9.37 m) depths were found. There was a tendency for more shallow

water tables to occur near the coast and deeper water tables inland in both sampling times.

Bores <15 m sampled in January 2004

nested bores sampled in January 2004

Bores >15 m sampled in January 2004

7.57.8

5.9 10

4.65.85.2

3.7 6.3

9.42.3

5.5

6

3.93.8

7.12.8/6.3-7.7

3.4

5.76

5

1.9-3.1

2.2-3.6

6.3-7.3

1.0-1.9

2.8 3.7254.8

6.68.8

6.2

3

4.9

2.25 2

3.2

2.8

5.84.2

5

3.3

5

4.1

5

6.1

2

5.5

4.2

500000 505000 510000 515000 520000 525000 530000 535000 540000 545000 550000 555000 560000 5650007800000

7805000

7810000

7815000

7820000

7825000

7830000

7835000

7840000

7845000

7850000

Bores sampled in September/October 2003

Figure 3. Depth to the water table during September/October 2003 and January 2004

The pH of the groundwater was mostly neutral to slightly alkaline, ranging from 5.7 to 7.5

with no spatial or vertical patterns. The groundwater EC varied from 0.1 to 119.4 mS/cm (EC

of sea water is 50 mS/cm). Groundwater with EC of less than 2 mS/cm generally occurs in

the inland at least to a depth of 50 m. EC showed an increasing trend with depth in the

nested bores except bore number 12000204 located in inland. The groundwater in the


11

nested bores (12100166) located near to the coast were mostly brackish reaching 15 mS/cm

at a depth of 9 m and 119 mS/cm at a depth of 51 m.

4.1.1.2 Major ions Major ions found in the groundwater of the lower Burdekin aquifer were calcium, magnesium,

sodium, bicarbonate and chloride. There do not appear to be any spatial or vertical patterns

related to cation and anion concentrations however, most of the cation and anion

concentrations are low between easting of 530000 and 540000 (Figure 4.1 & 4.2).

Figure 4. 1 Spatial and vertical distribution of major anions (in mg/L) in groundwater.

SO42- concentration.

77

742 431

8

4

10

224

1716 19220

303523 7

0102030405060

520000 525000 530000 535000 540000 545000 550000 555000

Eastings

Dep

th (m

)

Cl- Concentration

115

3013020 194

99

277

198

2930

5850 4151923

1783435186258 52

0102030405060

520000 525000 530000 535000 540000 545000 550000 555000

Eastings

Dep

th (m

)

HCO3 - Concentration

213

179264110 252

125

47

115

109 206152

117 2323169 5853223420596270 382

0102030405060

520000 525000 530000 535000 540000 545000 550000 555000

Eastings

Dep

th (m

)


12

Figure 4. 2 Spatial and vertical distribution of major cations (in mg/l) in the

groundwater.

In general, the groundwater in the nested bores near to the coast had high concentrations of

major ions compared to the bores inland. Also the concentrations of these major ions

showed an increasing trend with increasing depth except bore 12000204 which is located

Ca2+ Concentration

23

131913 89

12

45

29

2712

2938 1548 22

2021144025 22

0102030405060

520000 525000 530000 535000 540000 545000 550000

EastingsD

epth

(m)

Na+ Concentraion

166

4014022 257

79

84

106

2647

5534 9051519

246716162214 110

0102030405060

520000 525000 530000 535000 540000 545000 550000

Eastings

Dep

th (m

)

Mg2+ Concentration

12156 36

7

30

18

1324

1210 725 8

3013112618 16

21

0102030405060

520000 525000 530000 535000 540000 545000 550000

Eastings

Dept

h (m

)


13

more inland (Figure 5). This is consistent with the pattern of EC of these samples at different

depths. In bore 12000204 the major ion concentrations and the EC were greater in shallow

pipes compared to the deeper ones. There was no obvious vertical trend emerged in the

distribution of the major ions from the individual bore measurements.

Figure 5. Concentration of major cations and anions as a function of depth in the nested bores.

Calcium concentration of groundwater sampled from lower Burdekin aquifer range from 0.9

to 3171 mg/l, with concentrations less than 45 mg/l occuring down to a depth of 50 m in

inland. Magnesium concentrations range from 0.4 to 5747 mg/l with concentration exceeding

37 mg/l occurs mostly along the coast. Sodium concentration in the groundwater was the

greatest among the cations and range from 4.8 to 22909 mg/l, however concentration did not

exceed 260 mg/l in inland atleast to a depth of 50 m. Sodium concentration in the sea water

is about 10,000 mg/L (Todd, 1980).

Chloride concentrations of groundwater sampled from lower Burdekin aquifer range from 3.6

to 54000 mg/l (sea water contains 19,300 mg/L). The lowest concentrations, less than 280

Ca2+

0

10

20

30

40

50

60

70

80

90

100

0.00 500.00 1000.00 1500.00 2000.00 2500.00 3000.00 3500.00

Conc. (mg/l)

Dep

th (m

)

Bore 166

Bore 204

Bore 263

Bore 257

Bore 268

Mg2+

0102030405060708090

100

0.00 1000.00 2000.00 3000.00 4000.00 5000.00 6000.00 7000.00

Conc. (mg/l)

Dep

th (m

)

Bore 166

Bore 204

Bore 263

Bore 257

Bore 268

Na+

0102030405060708090

100

0 5000 10000 15000 20000 25000

Conc. (mg/l)

Dep

th (m

)

Bore 166

Bore 204

Bore 263

Bore 257

Bore 268

Cl-

0

10

20

30

40

50

60

70

80

90

100

0 10000 20000 30000 40000 50000 60000

Conc. (mg/l)

Dep

th (m

)

Bore 166

Bore 204

Bore 263

Bore 257

Bore 268

SO4 2-

0102030405060708090

100

0 200 400 600 800 1000 1200 1400

Conc. (mg/L)

Dep

th (m

)

Bore 166

Bore 204

Bore 263

Bore 257

Bore 268

HCO3-

0102030405060708090

100

0.0 100.0 200.0 300.0 400.0 500.0

Conc. (mg/l)

Dep

th (m

)

Bore 166

Bore 204

Bore 263

Bore 257

Bore 268


14

mg/l occur primarily in the inland areas down to a depth of 50 m. Chloride concentrations

exceeding 14000 mg/l occur near to the coast even at a depth of 18 m. Bicarbonate

concentrations range from 31 to 532 mg/l. Sulfate concentrations vary from 1 to 1188 mg/l in

the groundwater of the lower Burdekin aquifer however, concentrations were less than 80

mg/l in inland.

4.1.1.3 Nitrate

shallow bores <15 metres

nested bores

deep bores >15 metres

0.52

0.20

0.26 0.04

0.480.11

6.13

0.00/0.560.05

8.845.82

7.40

2.28

0.49

0.05

2.45

3.99/0.00

1.93

0.07

0.51

7.18

0.00-0.14

0.00-0.10

0.00

0.00-8.00

520000 525000 530000 535000 540000 545000 550000 555000 5600007815000

7820000

7825000

7830000

7835000

7840000

7845000

Figure 6. Spatial distribution of NO3-N in the lower Burdekin area during January 2004 The overall distribution of nitrate in the groundwater during the January 2004 is presented in

Figure 6. Although the nitrate concentrations in the majority of the bores were below the

ANZECC levels for long term environmental sustainability (5 mg NO3--N/L) values as high as

14.4mg of NO3--N/L were measured. During the September/October 2003 sampling period,

only 7 of the 30 bores showed NO3--N levels higher than 0.1 mg/L, with four bores above 5

mg/L. In contrast, 5 (23%) of the single bores yielded concentrations as low as 0.1 mg/L in

the January 2004 sampling period. There are five single bores which have NO3--N values

above 5 mg/L. Elevated nitrate levels were mostly observed in the Airville-Homehill area.

This applied to both sampling periods. Brodie et al (1984) and Weier (1999) also reported a

similar spatial distribution pattern but the magnitude of the concentrations they reported were

higher.


15

The vertical distribution in nitrate concentration revealed different trends between sampling

periods. In the first sampling period no nitrate was found at shallow depths (< 30 m) while

values as high as 8.8 mg NO3-N /L with a median of 0.52 mg/L were found in the second

sampling (Figure 7). The observed differences in the vertical nitrate concentration profile with

time could be related to fertiliser applications and the subsequent irrrigations, and recharge

through the soil. It is worth noting that the bores sampled during January 2004 were not the

same as those sampled in September/October 2003. A steady decline in nitrate

concentrations from the wet season to the dry season has been noted previously by Brodie

et al. (1984). Also, increases in concentrations have been found in other shallow unconfined

aquifers when recharge occurs (Littke and Hallberg 1991). Elevated nitrate concentrations

occurred mostly in bores located in areas where water table depths ranged between 2 and 6

m. Similar patterns have been observed with nitrate levels highest near the water table and

decreasing to below detection with depth both at this site by Brodie et al. (1984) and at other

sites by Trudell et al. (1986) and McLarin et al. (1999).

NO3-N (mg/L)

0 2 4 6 8 10 12 14 16

Dep

th (m

)

20

40

60

80

100

During January 2004During September/October 2003

Figure 7. NO3-N concentrations at different depths

With the exception of one shallow pipe located in the Ayr, none of the nested bores had

concentrations above 0.1 mg/L irrespective of the sampling time. A very shallow water table

(0.9 m) in the area could be the reason for the high nitrate concentration in one of the pipes.

Due to undetectable nitrate levels, no vertical variability was evident from the nested bores.

The nested bores were located along the coast. The nitrate concentrations in the farmer

production bores varied from 0 to 11 mg NO3--N/L and these are mostly located in the Ayr

area. Nitrite concentrations were below detection or negligible (<0.1 mg/L) in all of the bore


16

samples. Concentrations of NH4+ reached 8.1 mg/L in the September/October sampling

period while lower concentrations (maximum of 3.6 mg/L) were measured in bores sampled

during January 2004. Its distribution showed no particular spatial or vertical trend however,

nested bores showed higher values.

4.1.1.4 Ferrous iron The distribution of ferrous iron in the aquifer during the 2 sampling periods is shown in Figure

8 with values ranging from 0 to 361 mg/L. Kelly (1969) reported ferrous concentrations as

high as 100 g/m-3 in the Lower Burdekin aquifer. There was no evidence of a typical temporal

pattern. The bores sampled during September/October 2003 had very high concentrations

which might be related to the particular location of the bores.

While we have only limited data at this stage, there is some evidence of spatial trends with

high ferrous concentrations in the Kalamia Mill and Ayr areas. These results are consistent

with the findings of Kelly (1969). He suggests the nature of the Burdekin alluvial deposits

resulting in a random distribution of the ferrous levels. Elevated ferrous levels were also

observed in Giru-Barratta area where concentrations of 3.0 - 15.4 mg/L were measured.

Ferrous levels were below 2 mg/L in Airville - Home Hill area, and it is worth noting that

elevated nitrate concentrations were observed in these areas. In general, elevated nitrate

concentrations were not found in bores along the coast where elevated ferrous

concentrations were common.

As shown in Figure 9 ferrous iron concentrations were quite variable but tended to be

higher in bores less than 20 metres deep, compared to deep bores (>55 metres deep)

where the concentration ranged between 1.2 and 3.2 mg/L. The spatial and vertical trends

in ferrous iron concentrations were consistent in both the sampling periods.

Five nested bores along the coast were sampled to assess the vertical distribution in

ferrous iron in January 2004. Ferrous iron concentration in the nested bores ranged from

1.2 to 13.5 mg/L, and with the exception of one bore, there was no distinct vertical pattern.

This could be influenced by the fractured nature of the aquifer materials and the prevailing

geochemical conditions. Elevated ferrous concentrations in the groundwater suggest that

the aquifer has a suitable redox environment for denitrification. Ferrous levels and their

relationship with the geochemical conditions are discussed further in the following sections.


17

87

361160-2

42

1

1133

0

0

19

47

1-9

0 0 01

0

0

00

27

00

0

510000 515000 520000 525000 530000 535000 540000 545000 550000 555000 560000Easting

7814000

7819000

7824000

7829000

7834000

7839000

7844000

7849000

Nor

thin

g


nested boresdeep bores >15 metres

1.14.5

1.1 4.9

14.83.0

1.3

0.88/0.79 13.4

1.51.4

3.8

8.3

2.8

1.3

1.9

3.25/1.22-3.36

1.4

15.4

0.7

1.9

1.69-13.54

1.31-3.20

2.12-6.95

1.15-3.63

520000 525000 530000 535000 540000 545000 550000 555000 5600007815000

7820000

7825000

7830000

7835000

7840000

7845000

Figure 8. Distribution of ferrous iron in the lower Burdekin groundwater a) during September/October 2003 b) January 2004

Easting


18

Ferrous iron concentration (mg/L)

0 4 8 12 16 20 100 200 300D

epth

(m)

20

40

60

80

100

During January 2004 During September/October 2003

Figure 9. Ferrous levels at different depths during September/October 2003 and in January 2004.

4.1.1.5 Dissolved Organic carbon (DOC) DOC has been reported as the most thermodynamically preferred electron donor for

denitrification to occur (Spalding and Parrott 1994). In general DOC levels are low in

groundwater (1-5 mg/L) (Starr and Gillham 1993) with the highest value we found reported

in the literature being 27 mg/L in a shallow aquifer of a coastal plain (Spalding and Parrott

1994). The DOC levels in the Lower Burdekin aquifer ranged from as low as 4 mg/L to 82

mg/L (Figure 10) which is very high compared to the values reported in other aquifers. One

reason for these high levels could be the movement of sugar juices lost during harvest

moving to deeper depths and entering the shallow groundwater systems. These high DOC

concentrations in groundwater are also not that surprising when we compare them with the

high DOC concentrations of 260 mg/L that have been measured in the first runoff water from

one of the sugar cane farms. Assuming 30-40% of the irrigation is lost as deep drainage 78-

104 mg/L of DOC could be expected in the groundwater if no retention reactions occurred in

the unsaturated zone. A conservative estimation of Rayment in 2000 reported 0.14 tonnes of

sucrose-C/ha/yr (0.33 tonnes of sucrose/ha/yr) loss to the soil/water environment during

cane harvest.

While there were no specific spatial or vertical trends in DOC, when plotting DOC

concentrations against depth, there was a tendency for the highest DOC values to occur over

a smaller depth range compared to the low DOC values. This result must however be treated


19

cautiously since the data may be biased by the small number of deep bore samples obtained

in this study. DOC in the 5 nested bores ranged from 17 mg/L to 58 mg/L with no specific

patterns emerging.

In general DOC is inversely related to the depth of the water table. In places where water

tables are deeper, most of the DOC from the soil horizon is retained by sorption or consumed

by micro organisms in the unsaturated zone before reaching the water table. There is no

relationship between DOC and water table depth found in this study.


nested bores

deep bores >15 metres

46

31

52 20

3926

14

24/2124

56

16

27

51

5

17

12

82/17-47

56

55

45

48

20-48

32-58

28-31

32-50

520000 525000 530000 535000 540000 545000 550000 555000 5600007815000

7820000

7825000

7830000

7835000

7840000

7845000

Figure 10. Spatial distribution of DOC during January 2004

4.2 Geochemical conditions In most bores, particularly nested, dissolved oxygen (DO) concentrations were below 1 mg/L

in both sampling periods. DO ranged from 0.07 to 4.25 mg/L with the exception of one bore

that showed value as high as 6.4 mg/L. This range is in agreement with the work carried out

by Kelly (1969) in the lower Burdekin. Although the concentration varied considerably at

shallow depths it decreased with the water table depth. The cause for the decrease in DO

concentration with depth suggests that the oxygen was being used within the system to

oxidise reduced materials. In the nested bores DO concentrations did not show the

decreasing trend with depth as reported by others (Kelly, 1999). They were however, all

below 0.7 mg/L meaning that reduced conditions prevailed in the groundwater even at

depths of 88 m.


20

The most important factor determining the speciation and concentration of nutrients in

groundwater is the redox potential of the aquifer especially the presence of oxygen. Redox

potential (Eh) is measured directly in the field using platinum combined with hydrogen

electrodes. Many investigators have doubts about considering the Eh data quantitatively and

also report that they have little thermodynamic value as it is difficult to measure and interpret

(Postma et al. 1991; Spalding and Parrott 1994). Therefore, redox conditions are better

described by redox zones that are identified by redox species rather than the Eh

measurements themselves, which are rather used as supporting evidence. For example,

oxygen concentration will be close to saturation in the oxic zone, while the reduced zone is

likely to be characterised by elevated ferrous concentrations. Where substantial levels of

dissolved ferrous and manganese (5-10 g/m3) exist it is reasonable to assume that redox

equilibria condition will be well within the reduced range. Both metals tend to form insoluble

oxyhydroxide solids under oxidising conditions (Stumm and Morgan 1981).

Our data showed that Eh varied between -120 and 235 mV, with little evidence of trends with

increasing depth. More negative values of Eh however, were observed at shallow depths. As

a rule of thumb, high levels of oxygen saturation should correspond to relatively high values

of Eh (Kelly, 1969). In the bores sampled, DO varied from 0.1 to 6 mg/L between an Eh of 0

and 235 mV but was <1 mg/L below 0 mV (Figure 11).

Figure 11. Relationship between Eh and the redox species

D O concentra tion (m g/L)

0.5 1.0 1.5 2.0 2.5 3.0 3.5 4.0 4.5 5.0 5.5 6.0 6.5 7.0

Eh

(mV

)

-150

-100

-50

0

50

100

150

200

250

N O 3-N concentration (m g/L)

0.5 1.0 1.5 2.0 2.5 3.0 3.5 4.0 4.5 5.0 5.5 6.0 6.5 7.0 7.5 8.0 8.5

Eh

(mV)

-150

-100

-50

0

50

100

150

200

250

M anganese concentration (m g/L)

2 4 6 8 10 12 14

Eh (m

V)

-150

-100

-50

0

50

100

150

200

250

F e rro u s iro n co n ce n tra tio n (m g /L )

2 4 6 8 1 0 1 2 1 4 16 18

Eh

(mV)

-15 0

-10 0

-5 0

0

5 0

10 0

15 0

20 0

25 0


21

Elevated NO3--N concentrations occurred in groundwater where measured Eh were between

37 and 235mV, and NO3--N were largely undetectable in groundwater with Eh less than 0.

Nested bores were excluded in this comparison as all (except one pipe) had nitrate levels

below 0.1 mg/L. Total dissolved manganese concentrations ranged from 0 to 12 mg/L.

Manganese concentrations exceeding 3 mg/L occurred in groundwater at Eh between +50

and +110mV. Concentrations between 0 and 3 mg/L occurred in the Eh of 0 to -120 mV and

undetectable amounts were measured when Eh values exceeded +110mV. With the

exception of 2 bores manganese was undetectable in groundwater with DO values above 0.7

mg/L. In contrast to the NO3--N profile, elevated ferrous iron concentrations (>4 mg/L)

occurred in groundwater with Eh below 0 mV and a DO concentration lower than 1 mg/L.

Puckett and Cowdery (2002) reported ferrous levels of 135 µmols/L in the groundwater and

claimed such high concentrations are the result of ferric oxidation as the predominant

terminal electron accepting process. This could be true as well in our study as the Eh was

very low. Elevated ferrous iron concentrations were measured in negative Eh levels (below

0 mV) with concentrations in the range of 0.7-3.8 mg/L when the measured Eh was between

0 and 150mV. A comparison of ferrous concentrations at higher Eh values is not possible in

this study as groundwater in this area measured a maximum of +235 mV. The relationships

between oxygen, nitrate, manganese, ferrous species and the Eh levels characterise to

some extent the thermodynamics of the groundwater chemistry. The presence of ferrous iron

in the water samples indicates that water is suboxic.


22

5 Discussion 5.1 Major ion chemistry In general, the cation and anion concentrations in the groundwater appear to define three

zones (Figure 4.1). Groundwater in the zone 2 (Eastings from 530000 and 540000) had

lower concentrations of cations and anions compared to zone 1 (Eastings from 520000 and

530000) and zone 3 (Eastings from 540000 and 550000). The contrast between the major

cation (Ca, Mg ) and anion (Cl) concentrations in the zone 1 and zone 2 is likely the result of

the high irrigation rate in zone 2 (20 ML/ha/yr) compared to the zone 1 and 3 (10 ML/ha/yr).

Since the Cl, Ca and Mg concentrations in zone 1 and 3 are very similar, these cation and

anions are most likely the result of agrochemical inputs from fields. Another possible

explanation for the elevated concentrations in zone 3 might be the mixing of the water in

zone 3 with the intruded salt water.

Inorganic agricultural fertiliser is thought to be the major source of both nitrate and chloride in

the groundwater at the lower Burdekin, and the presence of chloride at all depths in the

inland suggests that fertiliser affected groundwater is present throughout the section.

Considering chloride a conservative tracer, differences in chloride concentrations could be

due to the spatial and temporal variability in infiltration. Chloride can be used as an indicator

to differentiate between the mixing of differently nitrate concentrated groundwater and the

biogeochemical processes (Min et al., 2003). Figure 12 shows the relationship between the

Cl- and NO3- in groundwater sampled mostly in the inland (to avoid very high concentrations

of Cl- from sea water intrusion) to a depth of less than 40 m. Points along the y-axis suggest

the presence of geochemical process attenuating nitrate. If there was no attenuation of

nitrate linearity behaviour between the two would be expected.

0

1

2

3

4

5

6

7

8

9

0 0.1 0.2 0.3 0.4 0.5 0.6 0.7

NO3- (mmol/L)

Cl- (m

mol

/L)

Figure 12. A plot of NO3- versus Cl- concentration in groundwater


23

The elevated bicarbonate concentrations in the lower Burdekin groundwater may reflect

some amounts of nitrate lost by denitrification. However, reactions such as the sulphate or

iron reduction also produce bicarbonate under anaerobic conditions. Very small amount of

sulphate found in inland could be an indication of sulphate reduction in the aquifer.

5.2 Redox environment of the aquifer The relationships between oxygen, nitrate, manganese, ferrous iron and the Eh levels

characterise to some extent the thermodynamics of the groundwater chemistry. The sharp

decline in DO and nitrate concentrations at an Eh of +100 can be interpreted as a redox

sequence such as that described by Vance, (2002), with oxygen consumption followed by

denitrification. Also the manganese and ferrous iron profiles patterns agree with the reported

redox sequence where manganese oxide reduction occurs after denitrification and before the

appearance of ferrous (Puckett and Cowdry 2002). There was however, no clear or sharp

boundaries separating the various zones.

As described in Section 2 aerobic degradation of the organic carbon is the most preferred

reaction of microbes to obtain energy. Therefore, it is possible that high loading of DOC into

the lower Burdekin groundwater consumes available oxygen which then results in an

anaerobic condition. Measured DO levels of mostly less than 2 mg/L in the samples suggests

the aquifer is in anoxic environment. Change in the redox conditions affects the chemical

composition of groundwater, mobility of ferrous or ferric iron and also the biodegradation of

organic contaminants such as pesticides. Furthermore, it is likely that the reduced

environment promotes sulphate reduction and methane fermentation at least in parts of

aquifers.

5.3 Nitrate reducing processes Nitrate is a redox-sensitive species and it would be expected that loss of nitrate due to

reduction or denitrification will be accompanied by changes in other redox-active species,

changes in geochemical conditions, and the production of reduced forms of nitrogen. As

such, an evaluation of the dominant species of the major redox-active elements including

oxygen, iron, manganese, sulphur, dissolved organic carbon, and nitrogen together with the

geochemical conditions can provide clues to the potential of the aquifer for denitrification

(Puckett and Cowdery 2002). For denitrification processes to happen groundwater oxygen

levels should be low. DO level of less than 2.0 mg/L was reported to be favourable for

denitrification (Korom 1992). Eh values less than < 250 mV has been reported suitable for

denitrification (Min et al, 2001). The DO concentrations and Eh measured in the lower

Burdekin groundwater are within the reported limits suitable for denitrification.


24

DOC has been referred to as the primary electron donor for denitrification processes and

also the most limiting electron donor in groundwater (Korom 1992; Spalding 1994). Adding a

carbon source to aquifers low in DOC has been chosen for in-situ nitrate remediation

purposes (Tompkins et al, 2001) to provide the substrate necessary for increasing biomass

and for providing the electron donor necessary for oxygen and nitrate reduction. In the lower

Burdekin groundwater however, DOC was much higher compared to the other aquifers in

general. Primary source of DOC in the lower Burdekin groundwater is the spilt sugars

during harvest. DOC from the spilt sugars surface can enter the groundwater through

recharge and/or mixing with surface water. Elevated levels of DOC in the groundwater

make the redox environment of the aquifer to be much reduced. This level of reduced

environment, in other words excess DOC enhances the weathering of the Fe and Mn

hydroxides to reduced forms of manganese and iron in the groundwater. While it may be

possible that this process occurs in the lower Burdekin aquifer which may have resulted in

the elevated levels of ferrous iron in the groundwater and Eh measurements as low as -120

mV, there is no clear evidence.

Korom (1992) stated that given the stoichiometry of 1:1.25 between nitrate and DOC on a

molar basis there must be 1.25 % more DOC than nitrate required for denitriifcation. Given

the range of DOC from 4 to 82 mg/L and the range of nitrate from 0 to 8.8 mg/L at a depth of

< 30 m, we should not expect to see nitrate at this depth (Figure 13). The fact that we do

suggest that either the biological and/or the thermodynamic constraints exist or, that the

denitrification rate is slow. When looking at the 4 points with high DOC levels and high nitrate

levels in Figure 13 it seems more likely that the biology or the type of DOC is limiting the

denitrification process than the DO levels. DO was less than 2 mg/L for these samples.

NO3-N concentration (mg/L)

0 2 4 6 8 10 12 14

DO

C c

once

ntra

tion

(mg/

L)

0

20

40

60

80

100

In January 2004

Figure 13 NO3-N concentration as a function of DOC


25

A plot of nitrate versus ferrous iron or ammonium in the groundwater shows the relationship

between these parameters; the presence of one is usually marked by the absence of other

(Figure 14). It appears that two processes involved, one relating to ferrous and the other to

high ammonium levels. As discussed in the literature bacteria uses iron in it’s reduced

ferrous form as an electron donor to reduce nitrate in the absence of dissolved oxygen. The

negative correlation between the nitrate, and the ferrous iron suggests that the ferrous iron in

the groundwater is used for reducing nitrate in the denitrification processes. While elevated

ferrous iron levels found in the lower Burdekin aquifer could be due to the pyrite weathering,

it is also possible that the ferrous iron starts to accumulate in the groundwater when both O2

and NO3- have been consumed so that micro organisms use ferric oxides as electron

acceptors and DOC as the electron donor. However, it is not clear at this stage to confirm

which process dominates.


0 2 4 6 8 10 12 14

Ferr

ous

iron

conc

entra

tion

(mg/

L)

0

10

20

30

40

50

During January 2004During September/October 2003

Figure 14. NO3--N concentration as a function of ferrous iron

Based on the stoichiometry of denitrification in equation (1), 10 moles of ferrous iron is

required for 2 moles of nitrate. At the median nitrate-N concentration of 0.51 mg/L at least

10.2 mg/L of ferrous iron is required. It appears that the ferrous iron available is not enough

for denitrification in the entire area of the lower Burdekin however, the higher amounts of

ferrous iron found in the Kalamia mill-Ayr area indicates that this area has good potential for

denitrification.

Based on a conservative assessment a map showing potential areas for denitrification in the

lower Burdekin was prepared with reference to the ferrous iron concentrations used by the

Minnesota Pollution Control Agency (1999) to define nitrate sensitive areas. Figure 15 shows

the potential areas for denitrification at shallow (a) and deep depths (b). The bores (eg.

112000114, 12000079 and 12000204) marked as high potential area but had elevated nitrate


26

levels because the water has higher DO or Eh which is not conducive to denitrification. The

map suggests that the shallow groundwater has potential for denitrification in most of the

areas compared to the deep groundwater.

Low potential for denitrification (with ferrous < 0.1 mg/L)

11910036

11910048

11910049

11910117

11910119

11910808

11910810

1200007912000204

11910038

119100561191015011910263A

12100012

11910051

11910264D

11910046

12100166F

1191020311910204

11910023

1190014911900150

500000 505000 510000 515000 520000 525000 530000 535000 540000 545000 550000 555000 560000 5650007800000

7805000

7810000

7815000

7820000

7825000

7830000

7835000

7840000

7845000

7850000

7855000

High potential for denitrification (with ferrous > 1 mg/L)

Bores with < 15 m

Low potential for denitrification (with ferrous <0.1mg/L)

1191006611910073

11910082 11910095

119101241191019011910744

11910942

1200011212000114

1210016612100166121001661210016612100166

1191026311910263119102631191026311910263

11910257119102571191025711910257

1200020412000204

119102681191026811910268119102681191026811910263E11910263D11910263B

11910842

119108861190016211910249

11910260

11911056

12100166E12100166D12100166C12100166B

1200009011910984

11910259

1190017911900131

11910270E11910270A

11900191

500000 505000 510000 515000 520000 525000 530000 535000 540000 545000 550000 555000 560000 5650007800000

7805000

7810000

7815000

7820000

7825000

7830000

7835000

7840000

7845000

7850000

High potential for denitrification (with ferrous >1 mg/L)

Bores with > 15 m

Figure 15. Potential area map for denitrification in the lower Burdekin


27

The existence of low nitrate concentrations in bores with high levels of ammonium, suggest

that Dissimilatory Nitrate Reduction (DNRA) could be occurring (Figure 16). Ammonium

(NH4+) is the end product of DNRA, which is a process that can temporarily remove nitrate as

long as NH4+ does not come in contact with aerobic environments (Smith et al. 1991; Korom

1992; Tesoriero et al. 2000). These results are also consistent with those reported by

Lamontagne et al. (2003) in an alluvial aquifer.


0 2 4 6 8 10 12 14 16

NH

4+ -N c

once

ntra

tion

(mg/

L)

0

1

2

3

4

5

During September/October 2003During January 2004

Figure 16 NO3--N concentration as a function of NH4

+-N

EC (m

S/cm

)

0 2 4 6 8 10

NH4+-N concentration (mg/L)

0

20

40

60

80

100

120

140

During January 2004During Seeptember/October 2003

Figure 17 NH4+-N vs EC

Figure 17 shows there was a close relationship between the EC and the ammonium

concentrations. Bores with EC higher than 20 mS/cm showed elevated ammonium levels

and these bores are located along the coast. Dissimilatory reduction of nitrate to ammonium


28

appears more in salt marshes and in strongly anoxic marine sediments (Tiedje 1988; Postma

et al. 1991). Marine deposits are present within the Burdekin aquifer which seems to support

the existence of dissimilatory denitrification processes. Furthermore, Tiedje et al. (1982)

stressed the competition between DNRA and denitrification which results in nitrate loss in a

system. He hypothesized that DNRA is favoured when nitrate is limiting and denitrification is

favoured when DOC is limiting. It appears that the DOC is not a limiting factor in the Burdekin

aquifer which tends to support the occurrence of DNRA. Also the greater flux of organic

carbon into the saturated zone produces high C:N ratio, which could promote the

transformation of nitrate to ammonium instead of N2 (Starr and Gillham, 1993). In the lower

Burdekin groundwater, elevated ammonium levels were found with DOC concentrations from

28 to 58 mg/L. Ammonium nitrogen however, can be formed from the mineralisation of

organic nitrogen (McMahon., 1999) in aerobic conditions.

Although nitrate is always targeted as the nitrogen species discharging into the ocean

ammonium and dissolved organic nitrogen (DON) can also be discharged to the ocean.


29

6 Conclusion Given the spatial variation in nitrate concentrations and the fact that the nitrate input is not

constant with time, it is difficult to determine the extent of nitrate reduction which has taken

place within the aquifer with the current level of data. However, the data on DO and Eh are

within the upper threshold levels reported in the literature, which would support the

occurrence of denitrification. Furthermore, elevated concentrations of ferrous iron and

ammonium, and their negative correlation with nitrate are consistent with denitrification

processes and dissimilatory nitrate reduction to ammonium as the causes for the reduction in

nitrate in the lower Burdekin groundwater, at least in some areas. Higher levels of ferrous

iron found both in the Ayr-Kalamia Mill area and near the coast indicate that these areas

have greater potential for removing nitrate from the groundwater.

A stoichiometry of 1:1.25 between NO3 and DOC for nitrate reduction, and the very high

levels of DOC with a wide spatial distribution indicates that the whole area has the potential

for denitrification. However, excess supply of labile DOC and the anoxic environment it may

create in the aquifer encourages weathering rates of iron and manganese hydroxides.

Elevated metal (Fe and Mn) concentration in groundwater is an environmental issue and

could be more problematic than nitrate if used for irrigation. There are also implications

concerning the environment if this water discharges into streams or other waterways.

The lack of nitrate in the nested bores located along the coast further indicate that nitrate

reduction processes take place as the groundwater makes its way to the coast. However,

further investigation is required to differentiate the nitrate reduction processes and to

delineate nitrate insensitive or potential denitrification depths. This will help to decide the

depths of water extraction for different purposes.

The map presented in this report is a first step towards identifying those areas of the lower

Burdekin with most potential for denitrification. Although there is still much work to be done in

this area, the current findings are a valuable start in helping to determine the fate of nitrate

that enters the groundwater. Such work is important in helping understand the complexities

of the floodplain and in enabling resource managers to better target their management

practices.


30

7 Recommendations

Future work addressing the potential for the lower Burdekin aquifer to consume nitrogen through denitrification should involve:

1. Strategic groundwater monitoring to assess the seasonal and long-term changes in nitrate

2. Development of greater understanding of the geochemical conditions and the electron donor availability along the groundwater flow path at different depths. This could be used to demarcate nitrate sensitive zones and depths within the aquifer

3. Measuring the rate of denitrification and determining the rate limiting factors (DOC, Nitrate, Ammonium, DO etc)

4. Developing improved understanding of the geochemistry and biology of the aquifer to allow quantification of denitrification in the system

5. Determining the concentration of dissolved organic nitrogen (DON) in the groundwater

6. Determining the lability of DOC in groundwater and the long-term environmental risks associated with a large loading of DOC in the aquifer.


31

8 References Arunakumaren, N. J., McMahon, G.A., and Bajracharya, K. (2000) Water management in the Lower

Burdekin: Groundwater model conceptualisation. Department of Natural Resources, Brisbane. Barton, L., McLay, C.D.A., Schipper, L.A., and Smith, C.T. (1999) Annual denitrification rates in

agricultural and forest soils: a review. Australian Journal of Soil Research, 37, 1073-1093.

Böhlke, J.K., and Denver, J.M. (1995) Combined use of ground water dating, chemical, and isotopic analyses to resolve the history and fate of nitrate contamination in two agricultural watersheds, Atlantic coastal plain, Maryland. Water Resources Research, 31, 2319-2339.

Böhlke, J.K., Wanty, R., Tuttle, M., Delin, G., and Landon, M. (2002) Denitrification in the recharge area and discharge area of a transient agricultural nitrate plume in a glacial outwash sand aquifer, Minnesota. Water Resources Research, 38(7), 10.1029/2001WR000663, 200238, p. 10.1-10.26.

Bristow, K.L., and Charlesworth, P.B. (2002). Effective water management vital to the long-term economic viability of the Lower Burdekin In: Proceedings of the ANCID Conference, 1st -3rd September, Griffith, NSW, Australia.

Brodie, J.E., Hicks, W.S., Richards, G.N., and Thomas, F.G. (1984) Residues related to agricultural chemicals in the groundwaters of Burdekin river delta, North Queensland. Environmental Pollution, 8(7), 187-215

Charlesworth, P. B., and Bristow, KL. (2002) Sustainable Management of the Burdekin Groundwater System, Milestone report to the National program for Irrigation Research and Development.

Korom, S.F. (1992) Natural denitrification in the saturated zone: A review. Water Resources Research, 28, 1657-1668

Kolle, W., Strebel, O., and Bottcher, J. (1985) Formation of sulfate by microbial denitrification in a reducing aquifer. Water Supply, 3, 35-40

Kelly, G.J. (1974) Iron in Burdekin irrigation waters, Proc.Qld Soc. Sugar Cane Technol., 41, 27-35.

Kelly, W.R., (1997) Heterogeneities in ground-water geochemistry in a sand aquifer beneath an Irrigation field, Journal of Hydrology 198, 154-176, U.S.A.

Kelly, W.R., and Ray, C. (1997) Impact of irrigation on the dynamics of nitrate movement in a shallow sand aquifer, Final Report. Illinois State Water Survey, Champaign, IL.

Lamontagne, S., Herczeg, A.L., Dighton, J.C., Pritchard, J.L., Jiwan, J.S., and Ullman, W.L. (2003) Groundwater–surface water interactions between streams and alluvial aquifers: Results from the Wollombi Brook (NSW) study (Part II – Biogeochemical processes)’, CSIRO Land and Water Technical Report 42/03, 2003

McLarin, W., Bekesi, G., Brown, L., and McConchie, J. (1999) Nitrate contamination of the unconfined aquifer, Manakau, Horowhenua, New Zealand, Journal of Hydrology (NZ) 38(2): 211-235, s.l,

Minnesota Pollution Control Agency, (1999) Estimating ground water sensitivity to nitrate contamination’, Environmental outcomes division Ground water monitoring & assessment program.

Min, J., Yun, S., Kim, K., Kim, H and Kim, D. (2003) Geologic controls on the chemical behaviour of nitrate in riverside alluvial aquifers, Korea. Hydrological Processes. 17, 1197-1211.

McMahon, G.A., Arunakumaren, N.J., and Bajracharya, K. (2002) Estimation of the groundwater budget of the Burdekin River Delta aquifer, North Queensland. Department of Natural Resources and Mines, 80 Meiers Rd., Indooroopilly, Qld., Australia. 4075. Contributed paper to the International Association of Hydrogeologists Conference, Darwin, 14-17 May 2002

Narayan, K.A., Schleeberger, C., Charlesworth, P.B., and Bristow, K.L. (2003) Effects of groundwater pumping on saltwater intrusion in the Lower Burdekin delta, North


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Queensland. In: Proceedings of the MODSIM 2003 Conference, 14th – 17th July, Townsville, Australia

Postma, D., Boesen, C., Kristiansen, H., and Larsen, F. (1991) Nitrate reduction in an unconfined aquifer: Water chemistry, reduction processes, and geochemical modeling. Water Resources Research, 27, 2027-2045

Puckett, L.J., and Cowdery, T.K. (2002) Transport and fate of nitrate in a glacial outwash aquifer in relation to ground water age, land use practices, and redox processes. Journal of Environmental Quality, 31(3), 782-796

Rayment, G.E., and Higginson, F.R. (1992) Australian Laboratory handbook of soil and water chemical methods. Inkata Press. Melbourne.

Smith, R.L., Howes, B.L., and Duff, J.H. (1991) Denitrification in nitrate contaminated groundwater: occurrence in steep vertical geochemical gradients. Geochimica et Cosmochimica Acta, 55, 1815-1825.

Spalding, R.F., and Parrott, J.D. (1994) Shallow groundwater denitrification. The Science of the Total Environment, 141, 17-25

Spruill, T.B., Eimers, J.L., and Morey A.E. (1997) Nitrate-Nitrogen concentrations in Shallow ground water of the coastal plain of the Albemarle-Pamlico Draingae study unit, North Carolina and Virginia. Factsheet 241-96. United States Geol. Survey, Reston, VA.

Starr, R.C., and Gillham, R.W. (1993) Denitrification and organic carbon availability in two aquifers. Ground water, 31(6), 934-947

Stumm, W., and Morgan, J.J. (1981) Aquatic Chemistry, 2nd ed., 780pp., John Wiley New York.

Tesoriero, A.J., Liebscher, H., and Cox, S.E. (2000) Mechanism and rate of denitrification in an agricultural watershed: Electron and mass balance along groundwater flow paths. Water Resources Research, 36(6), 1545-1559

Thorburn, P.J., Biggs, J.S., Weier, K.L., Keating, B.A. (2003) Nitrate in groundwaters of intensive agricultural areas in coastal Northeastern Australia. Agriculture, Ecosystems and Environment, 94, 49-58.

Tiedje, J.M. (1988) Ecology of denitrification and dissimilatory nitrate reduction to ammonium. In “Biology of anaerobic Microorganisms”, edited by Zehnder, A.J.B. A Wiley-Interscience Publication, John Wiley & Sons New York, 179-244

Tompkins, J.A., Smith, S.R., Cartmell, E. and Wheater, H.S. (2001) Insitu bioremediation is a viable oprion for denitrification of Chalk groundwaters. Quarterly Journal of Engineering Geology and Hydrology.34, 111-125

Trudell, M.R., R.W. Gillham, and Cherry, J.A. (1986) An in-situ study of the occurrence and rate of denitrification in a shallow unconfined sand aquifer. Journal of Hydrology, 83, 251–268.

Weier, K., (1999) The quality of groundwater beneath Australian sugarcane field. CSIRO Tropical Agriculture, St. Lucia, 1999.


33

9 Appendix 1. Data Summary

9.1 Appendix 1A. Data summary for the samples taken during September/October 2003

Code RN EASTING NORTHING Sample date DEPTH pH DO EC Ferrous NO3-N NH4-N WT depth

(M) (mg/l) (mS/cm) (mg/l) (mg/l) (mg/l) m

1 11910038 545118 7843045 25-Sep-03 10.0 6.573 x x 86.7 <0.04 0.3 5

2 11910056 547551 7839832 24-Okt-03 8.3 x x 5.59 361.3 <0.10 1.5

3 11910150 548739 7840635 25-Sep-03 10.5 6.585 x 3.64 16.2 <0.04 0.5 2

4E 11910263E 548727 7840630 25-Sep-03 26.9 6.95 x 12.970 0.000 <0.04 0.53 2.5

4D 11910263D 548727 7840630 25-Sep-03 41.6 6.386 x 42.300 1.952 <0.04 3 3

4C 11910263C 548727 7840630 25-Sep-03 54.5 6.241 x 63.700 0.421 <0.04 6.2 3.5

4B 11910263B 548727 7840630 25-Sep-03 77.5 6.126 x x 1.020 <0.04 6.5 3.5

4A 11910263A 548727 7840630 25-Sep-03 88.5 6.119 x 69.400 1.947 <0.04 7.1 4

5 12100012 549722 7815700 29-Sep-03 10.9 6.600 x 5.22 41.9 <0.04 0.6 4.2

6 11910842 553234 7829630 06-Oct-03 41.6 6.550 0.05 1.82 1.4 0.05 0.2 4.1

7 11910051 527608 7841938 06-Oct-03 8.4 6.720 0.05 2.05 11.1 0.09 0.5 2.2

8 11910886 522262 7840040 10-Oct-03 26.9 5.720 0.07 47.70 2.8 <0.04 0.7 5

9 11900162 518841 7840069 10-Oct-03 55.8 5.920 0.08 51.60 2.5 <0.04 0.9 5

10D 11910264D 547551 7839832 17-Oct-03 9.7 7.030 0.08 1.38 0.0 <0.04 0.2 3.32

10B 11910264B 547551 7839832 10-Oct-03 88.5 6.290 0.08 60.90 0.7 0.35 0.3

11 11910249 527586 7841970 15-Oct-03 46.5 5.940 0.12 82.90 0.0 <0.04 2.4 2.8

12 11900196 512207 7829860 15-Oct-03 18.5 6.560 1.48 1.05 1.20 0.1 8.78

13 11910260 553234 7829630 17-Oct-03 19.9 5.960 0.14 64.40 0.0 <0.04 1.6 5

14 11910046 540248 7837272 17-Oct-03 15.3 6.600 0.09 1.66 18.8 <0.04 2.2 4.86

15 11911056 522375 7840914 17-Oct-03 26.7 5.890 0.10 59.20 47.2 <0.04 1.6 1.99


34

Code RN EASTING NORTHING Sample date DEPTH pH DO EC Ferrous NO3-N NH4-N WT depth

(M) (mg/l) (mS/cm) (mg/l) (mg/l) (mg/l) m

16F 12100166F 555706 7817979 29-Sep-03 9.2 6.679 0.65 19.66 6.0 <0.04 2.4 2.5

16E 12100166E 555706 7817979 29-Sep-03 17.8 6.573 0.60 71.70 9.0 <0.04 3.2 2.5

16D 12100166D 555706 7817979 29-Sep-03 23.2 6.342 0.31 92.50 1.2 <0.04 1.9 2.75

16C 12100166C 555706 7817979 29-Sep-03 28.8 6.285 95.10 0.0 <0.04 1.7 3

16B 12100166B 555706 7817979 29-Sep-03 34.4 6.311 0.38 117.20 0.0 <0.04 0.7 3.25

16A 12100166A 555706 7817979 29-Sep-03 57.9 6.160 0.30 118.70 0.4 <0.04 0.1

17 11910203 513360 7822970 02-Oct-03 14.2 7.200 0.84 4.15 0.0 <0.04 0.3 3.2

18 11910204 517618 7822257 02-Oct-03 13.0 5.754 0.09 1.07 0.0 0.37 0.2 6

19 12000090 544416 7823282 02-Oct-03 62.4 7.293 0.05 1.12 0.0 5.10 0.0 5.5

20 11910984 523850 7821575 02-Oct-03 52.7 6.786 2.21 1.15 1.2 7.00 0.1 6.1

21 11910023 534760 7823700 02-Oct-03 11.9 7.356 3.00 1.20 0.0 14.40 0.1 6.2

22 11900149 512175 7842250 06-Oct-03 9.3 6.840 0.05 34.00 0.0 <0.04 8.1 2.2

23 11900045 513208 7842138 06-Oct-03 7.1 7.150 5.85 2.32 x x x 2.75

24 11900150 512295 7842245 10-Oct-03 9.2 6.770 0.07 20.70 8.2 <0.04 4.1 2

25 11910259 551889 7835052 15-Oct-03 30.6 6.010 0.10 59.40 0.1 <0.04 1.4 4.23

26 11910256 549972 7834847 15-Oct-03 26.9 5.940 0.13 61.30 0.4 <0.04 0.8 5.83

27 11900179 516330 7839970 15-Oct-03 47.2 6.120 3.90 34.60 2.3 <0.04 0.4 4.76

28 11900131 518533 7841225 15-Oct-03 28.4 5.830 0.14 78.10 6.6 <0.04 1.7 3.68

29E 11910270E 539097 7844014 17-Oct-03 54.5 6.400 0.08 1.28 0.0 0.27 0.2

29A 11910270A 539097 7844014 17-Oct-03 77.5 5.970 0.10 72.00 0.0 <0.04 2.4 5.04

30 11900191 511995 7832160 17-Oct-03 29.0 6.450 0.09 0.74 0.0 <0.04 0.1 6.64

x = not measured


35

9.2 Appendix 1B. Data summary for the samples taken during January 2004

Bore RN Northing Easting Code Watertable Bore depth pH Eh DO EC Fe2+ Conc. mg/L DOC

[m] [m] [mV] [mS/cm] [mg/l] NH4-N NO3-N mg/l

>15metres deep

11910066 538512 7834955 9 7.46 35.5 6.74 23 1.189 0.00 0.06 0.52 46 11910073 543005 7833122 5 7.81 22.5 6.2 6 0.4 0.403 4.54 0.06 0.200 31.00 11910082 531452 7830101 14 5.86 24.9 6.35 45 0.996 1.08 0.11 0.26 52 11910095 538680 7830032 11 10.48 17.8 6.5 -33 0.247 4.95 0.06 0.04 20 11910124 546635 7836128 20 4.59 17 6.7 -25 0.3 3.03 14.81 1.65 0.48 14.00 11910190 522201 7836000 22 5.78 51.2 6.63 -83 0.602 3.02 0.12 0.11 26.00 11910744 529880 7833960 16 5.15 23.6 6.1 235 1.155 1.31 0.08 6.13 10.00 11910877 528950 7837340 17 3.65 45 6.52 200 0.939 0.79 0.08 0.56 24.00 11910942 535220 7836619 10 6.35 19.4 6.37 -15.5 0.12 13.44 0.34 0.05 24.00 12000112 542848 7827695 1 9.37 24.4 6.87 0.72 0.751 1.46 0.26 8.84 56.00 12000114 544253 7826081 3 2.29 17.4 6.37 204 6.38 0.543 1.36 0.07 5.82 16.00

<15metres deep

11910036 536273 7842875 7 5.55 11 6 37 0.502 3.81 0.06 7.400 27.00 11910048 544963 7839313 6 5.95 13.1 6.8 -120 0.41 0.672 8.34 0.35 2.28 51.00 11910049 543378 7835234 4 3.94 12 6.29 21 0.33 0.069 2.84 0.05 0.49 4.00 11910117 534480 7833000 15 3.76 10.7 6.6 80 0.172 1.27 0.08 0.05 17.00 11910119 536424 7837691 8 7.08 15.4 5.95 205 0.36 1.89 0.05 2.45 12.00 11910258 540587 7839733 21 2.8 9.1 7.2 116 0.2 1.593 3.25 0.11 3.99 82.00 11910808 534640 7826700 12 3.39 9.6 6.6 114 3.72 0.553 1.39 0.29 1.93 56.00 11910810 532815 7829579 13 5.69 9.2 6.6 -107 0.693 15.42 0.22 0.07 55.00 11910878 528951 7837340 18 3.68 10 6.26 135 0.815 0.88 0.08 0.000 21.00 11910975 526497 7831527 19 6.03 10.2 6.25 145 1.468 0.71 0.06 0.51 45.00 12000079 545366 7824115 2 4.95 14.1 6.75 109 1.28 0.937 1.87 0.08 7.18 48.00


36

Nested bores 12100166 548727 7840630 166A 3.05 51.9 6.19 160 0.65 119.4 8.62 0.10 0.000 21.00

166B 2.95 34.4 6.19 154.2 0.7 118.3 8.13 0.10 0.000 20.00 166C 2.61 28.8 6.2 84.4 0.73 96 2.00 1.70 0.000 29.00 166D 2.56 23.2 6.18 49.2 0.6 94.4 1.69 1.65 0.14 30.00 166E 2.3 17.8 6.45 -28.8 0.51 70.3 11.37 3.38 0.000 39.00 549657 7832526 166F 1.93 9.2 6.2 -85.3 0.52 15.29 13.54 3.11 0.000 48.00

11910263 1A 3.62 88.5 6 70 69.5 3.20 2.32 0.10 32.00 1B 3.57 77.5 5.98 70 70 2.63 2.68 0.000 44.00 1C 3.22 54.5 6.09 75 64.4 1.31 3.60 0.000 58.00 548514 7823409 1D 2.7 41.6 6.28 81 43.5 1.63 2.15 0.000 42.00 1E 2.16 26.9 6.78 32 9.35 2.91 0.72 0.000 44.00

11910257 2A 7.24 70.6 6.06 109 0.63 54.6 2.12 0.73 0.000 30.00 2B 7.16 60.6 6 111 0.68 54.1 2.75 0.69 0.000 28.00 555706 7817979 2C 6.36 33.5 6.45 -3.9 0.6 13.45 6.95 0.17 0.000 28.00 2D 6.3 20.7 6.58 -53.9 0.49 2.31 5.92 0.16 0.000 31.00

12000204 0A 1.86 32 7.1 125 1.12 1.205 1.20 0.05 0.000 40.00 0B 1.85 21.5 6.81 -70.2 0.52 1.544 3.03 0.05 0.000 38.00 0C 1.86 15 6.59 -51.5 0.56 1.465 3.63 0.05 0.000 32.00 0D 0.96 5.4 6.66 82 0.59 2.53 1.15 0.05 8.00 50.00

11910268 540568 7839714 3A 7.75 74.3 6.51 96 0.2 51.3 2.03 1.66 0.000 47.00 3B 6.41 59.1 7.55 -65 0.07 3.18 1.22 0.27 0.000 28.00 3C 6.3 42.8 6.65 13 0.07 0.339 1.70 0.40 0.000 25.00 3D 7.32 29.2 6.65 80 0.07 0.261 3.36 0.30 0.000 17.00 3E 7.39 19.2 5.97 -6.1 0.07 0.268 1.61 0.32 0.000 22.00


37

9.3 Appendix 1C. Cation and anion data for the samples taken during January 2004

mg/L Cations mg/L Anions mg/L

Sample Id Mn NH4 NO2 NO3 Ca Mg K Na CO32- HCO3- Cl mg/L SO4 mg/L PO4 mg/L

1 0.1 0.26 <0.5 8.84 11.77 24.10 1.1 47.27 15 205.5 30 4 0.03 2 0.1 0.08 <0.5 7.18 21.84 16.28 2.5 110 0 382.2 51 7 0.07 3 0.1 0.07 <0.5 5.82 28.60 12.01 2.8 55 0 151.8 63 17 0.06 4 <0.1 <.06 <0.5 0.49 3.48 1.60 1.9 4.8 0 31.0 4 1 0.08 5 <0.1 <.06 <0.5 0.20 12.78 11.45 5.1 40 0 179.1 30 7 0.04 6 0.1 0.35 <0.5 2.28 14.74 6.54 8.7 90 0 231.6 41 19 0.24 7 <0.1 0.06 <0.5 7.40 37.88 10.32 3 34 0 117.3 52 16 0.03 8 <0.1 <.06 <0.5 2.45 22.00 7.86 1.6 19.4 0 57.7 21 20 0.05 9 <0.1 0.06 <0.5 0.52 23.10 20.91 4.9 166 6 213.0 116 77 0.05 10 0.5 0.34 <0.5 0.05 27.49 12.61 3.1 26 0 108.6 28 22 0.03 11 0.1 0.06 <0.5 0.04 13.40 6.02 2.3 22 0 110.5 20 2 0.03 12 0.2 0.29 <0.5 1.93 20.69 13.24 2.9 71 19 234.3 32 3 0.59 13 0.3 0.22 <0.5 0.07 14.03 11.18 2 61 0 205.7 35 5 0.04 14 <0.1 0.11 <0.5 0.26 19.15 15.16 3.3 140 9 263.6 132 4 0.04 15 0.3 0.08 <0.5 0.05 7.76 4.48 3.1 15.3 0 68.7 19 2 0.16 16 <0.1 0.08 <0.5 6.13 44.93 30.23 4.7 84 0 46.6 282 4 0.07 17 <0.1 0.08 <0.5 0.56 29.30 18.55 3.5 106 0 114.7 196 10 0.15 18 <0.1 0.08 <0.5 <0.1 40.71 26.09 3.3 62 0 96.3 183 2 0.11 19 <0.1 0.06 <0.5 0.51 24.89 17.98 1.9 214 18 270.0 259 3 0.22 20 1.6 1.65 <0.5 0.48 88.80 36.72 13.1 257 0 316.7 190 431 0.04 21 <0.1 0.11 <0.5 3.99 19.94 30.38 1.1 246 25 532.2 152 30 0.08 22 0.4 0.12 <0.5 0.11 11.58 7.34 2.9 79 0 124.6 103 8 0.25

166A 0.2 0.10 <0.5 <0.1 3027.20 5303.76 310 22744 0 356.3 52481 580 0.03 166B 0.3 0.10 <0.5 <0.1 3171.20 5747.04 321 22909 0 349.3 54002 447 0.03 166C 3.7 1.70 <0.5 <0.1 2852.40 5251.92 402 17746 0 285.9 43175 394 0.03 166D 3.7 1.65 <0.5 0.14 2436.80 3974.64 378 16730 0 266.1 41352 346 0.03 166E 6.2 3.38 <0.5 <0.1 684.00 2373.84 402 12339 0 435.9 27106 920 0.03 166F 2.9 3.11 <0.5 <0.1 242.74 348.38 47 3002 0 376.3 4282 814 0.03


38

0A <0.1 <.06 <0.5 <0.1 14.98 22.20 10.3 140.7 9 175.0 218 5 0.04 0B 0.8 <.06 <0.5 <0.1 33.59 36.00 7.4 207 12 226.5 297 8 0.03 0C 1.3 <.06 <0.5 <0.1 36.81 34.71 6.8 193 14 246.1 253 10 0.03 0D 0.1 <.06 <0.5 8.00 62.26 49.98 4.8 306.23 0 194.1 594 41 0.22 1A 5.9 2.32 <0.5 0.10 1260.12 1308.42 371 9191 0 379.6 27036 1165 0.04 1B 4.8 2.68 <0.5 <0.1 859.40 2231.04 363 12955 0 401.3 26271 1143 0.04 1C 12.1 3.60 <0.5 <0.1 661.80 1894.32 400 12114 0 431.7 22618 1188 0.06 1D 5.3 2.15 <0.5 <0.1 447.40 951.00 274 7748 0 378.9 14350 917 0.09 1E 0.6 0.72 <0.5 <0.1 55.51 101.52 55 1575 7 206.2 2609 149 2A 5.3 0.73 <0.5 <0.1 702.80 1764.84 285 9147 0 215.0 19703 670 2B 5.2 0.69 <0.5 <0.1 633.40 1620.12 273 8681 0 205.1 18375 668 2C 0.8 0.17 <0.5 <0.1 175.16 247.78 73 2254 0 140.7 4124 215 2D 0.4 0.16 <0.5 <0.1 16.99 29.52 15 292 0 138.2 448 22 3A 2.9 1.66 <0.5 <0.1 998.79 1021.44 283 7343 0 374.6 18040 781 3B 0.2 0.27 <0.5 <0.1 17.06 8.33 22 529 0 143.4 857 43 3C 0.1 0.40 <0.5 <0.1 1.19 1.98 5.7 70 0 117.7 33 8 3D 0.1 0.30 <0.5 <0.1 0.91 0.36 4.1 55 0 114.4 23 3 3E 0.1 0.32 <0.5 <0.1 1.05 0.70 4.5 49 0 118.5 24 3

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Land and Water - CSIRO - Assessment of the …Assessment of the geochemical environment in the lower...

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