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Page 1: Groundwater–surface water interactions between streams and … · 2005-10-14 · Groundwater–surface water interactions between streams and alluvial aquifers: Results from the

Groundwater–surface water interactions between streams and alluvial aquifers: Results from the Wollombi Brook (NSW) study (Part II – Biogeochemical processes)

Sébastien Lamontagne, Andrew L. Herczeg, John C. Dighton, Jodie L. Pritchard CSIRO Land & Water Jaswant S. Jiwan NSW Department of Infrastructure, Planning and Natural Resources William J. Ullman College of Marine Studies, University of Delaware

Technical Report 42/03, July 2003

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© 2003 CSIRO To the extent permitted by law, all rights are reserved and no part of this publication covered by copyright may be reproduced or copied in any form or by any means except with the written permission of CSIRO Land and Water.

Important Disclaimer

CSIRO Land and Water advises that the information contained in this publication comprises general statements based on scientific research. The reader is advised and needs to be aware that such information may be incomplete or unable to be used in any specific situation. No reliance or actions must therefore be made on that information without seeking prior expert professional, scientific and technical advice.

To the extent permitted by law, CSIRO Land and Water (including its employees and consultants) excludes all liability to any person for any consequences, including but not limited to all losses, damages, costs, expenses and any other compensation, arising directly or indirectly from using this publication (in part or in whole) and any information or material contained in it.

Cover Photograph

Other images available from CSIRO Land and Water Image Gallery: www.clw.csiro.au/ImageGallery/ Description: Fordwich site Photographer: Sébastien Lamontagne © 2003 CSIRO

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Summary The biogeochemistry of groundwater–surface water interactions in a subtropical sand-bed stream (Wollombi Brook, NSW) was studied over a two-year period using networks of piezometers and stream porewater profiles at three sites with different geomorphological properties. Several key findings emerged from the study: • While the lower section of the brook should be a discharge area for brackish regional

groundwater, the brook and the alluvial aquifer remained relatively fresh. Locally, dewatering of coal seams in active mines in the region has lowered the water table in the regional aquifer and reduced or reversed hydraulic gradients.

• Groundwater–surface water interactions were primarily controlled by lateral exchange processes between the brook and the alluvial aquifer, including bank recharge– discharge cycles during floods and hyporheic exchange during low flows.

• The alluvial aquifer generally had reducing conditions and tended to be enriched in chemical species more stable or more soluble in anoxic environments, including NH4

+, filterable reactive phosphorus (FRP), Fe2+, and Si.

• The redox status of the alluvial aquifer may be related to the depth to the water table in the floodplain. Reducing conditions were usually associated with shallower water tables.

• The presence of several redox interfaces throughout the alluvial aquifer probably fosters several cycles of nutrient transformations during a lateral exchange cycle. These would tend to promote the loss of some elements (such as nitrogen) relative to others (such as phosphorus).

• Under low flows, surface water is probably exchanged with streambed porewater every several hundred metres by hyporheic processes.

• Nutrient concentration in Wollombi Brook remained low throughout the study period. It is proposed that the large rates of lateral exchange with the sandy alluvial aquifer (especially hyporheic exchange) are fostering the maintenance of high water quality in the brook under low flows.

It is proposed that future research on the significance of groundwater–surface water interaction on nutrient cycles in streams and rivers should focus on: • The further development of measurement and modelling techniques to characterise

hyporheic exchange and associated biogeochemical transformations. • The integration of groundwater discharge and hyporheic exchange modelling to

provide more accurate estimates of nutrient fluxes from groundwater. • The use of hydrograph separation techniques to estimate the contribution of

groundwater discharge to stream nutrient budgets during flood pulses.

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Acknowledgements Many people in the Wollombi Brook catchment gave us enormous assistance and information during the course of our field studies. We especially thank Wendy and Bill Lawson, who were always helpful in supplying information, data and sometimes sustaining us with their first class semillon or chardonnay. Jim Maher enthusiastically collected water samples on a regular basis, often well beyond the call of duty when the brook was nearly dry, or during occasional flooding rains. The following people generously allowed us access to their properties to install piezometers and groundwater boreholes; Professor Tony Palfreeman at Wollombi, Mr. Andrew Furnance at Fordwich and Mrs. Heather Kannar at Warkworth. We acknowledge assistance provided by Peter Smith and Michael Williams from the NSW Department of Sustainable Natural Resources who helped get the project off the ground. This project was funded by Land and Water Australia, CSIRO Land and Water, and the NSW Department of Natural Resources.

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Table of Contents

RESULTS FROM THE WOLLOMBI BROOK (NSW) STUDY (PART II – BIOGEOCHEMICAL PROCESSES) ..................................................................................I

SUMMARY..........................................................................................................................I

ACKNOWLEDGEMENTS..................................................................................................II

TABLE OF CONTENTS ...................................................................................................III

INTRODUCTION................................................................................................................1

Literature Review ......................................................................................................... 2

Methods........................................................................................................................ 4

Site description............................................................................................................. 4

Hydrology ..................................................................................................................... 4

Riparian transects ........................................................................................................ 5

Sampling Techniques................................................................................................... 6

Field Trips..................................................................................................................... 9

RESULTS ..........................................................................................................................9

General patterns in water quality.................................................................................. 9

Spatial trends across the alluvial plain ....................................................................... 10

Patterns in nutrient concentration............................................................................... 13

Water Table profiles ................................................................................................... 17

Streambed porewater profiles .................................................................................... 18

Comparison of potential water fluxes from alluvial groundwater discharge and hyporheic exchange ................................................................................................... 23

Nitrogen mineralisation rates across the floodplain.................................................... 25

DISCUSSION...................................................................................................................26

Nutrient recycling in groundwater during lateral exchange......................................... 26

Dissolved nutrients in surface water........................................................................... 29

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Future Research......................................................................................................... 31

CONCLUSION.................................................................................................................34

REFERENCES.................................................................................................................35

APPENDIX 1. ANALYTICAL METHODS........................................................................40

APPENDIX 2. DATA SUMMARY ...................................................................................42

APPENDIX 3. SOME USEFUL TOOLS FOR GROUNDWATER–SURFACE WATER INTERACTION STUDIES................................................................................................56

Mini piezometer bundles ............................................................................................ 56

Drive Point .................................................................................................................. 59

Porewater profiler ....................................................................................................... 60

APPENDIX 4. PARSONS CREEK TRACER ADDITION ................................................61

Methods...................................................................................................................... 61

Results ....................................................................................................................... 62

Discussion .................................................................................................................. 64

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Introduction The risk of algal bloom formations brought by modifications to the flow regime of rivers and increased nutrient export from catchments is a continuing concern in Australia. While the emphasis in the past has been on nutrient export from surface runoff (especially P), it is now recognised that groundwater can also be involved in the transport of nutrients, especially nitrogen (N). For example, in Australia, groundwater has been found to be a significant source of N to some freshwater (Lamontagne 2002), estuarine (Linderfelt and Turner 2001) and coastal environments (Smith et al. 2003). The recognition that groundwater can contribute significantly to the nutrient load of some ecosystems will require some modifications to the approaches traditionally taken to manage nutrient export from catchments. For example, a common approach to assess nutrient inputs to surface water ecosystems at the regional scale is to use relationships between land-use and nutrient export in runoff (Young et al. 1996). This kind of approach may not be optimal for catchments where groundwater is a significant source of nutrients for two reasons. First, significant time lags (i.e. decades to centuries) can occur before groundwater systems reach a new equilibrium following a change in land use (Lamontagne 2002). Such time lags are well appreciated for some of the environmental problems related to groundwater, such as dryland salinity (Allison et al. 1990), but they are seldom considered in the case of nutrient discharge from groundwater. Thus, future nutrient inputs to many ecosystems are currently underestimated because the contribution from groundwater is yet to be fully realised. The second problem related to the estimation of nutrient discharge from groundwater is the poor understanding of the impact of groundwater–surface water interactions on the load and the form in which groundwater-borne nutrients are discharged to surface water. Of special interest is the role of the riparian zone in modulating nutrient inputs from groundwater (Hill 1996). As an interface (or ecotone) between terrestrial and aquatic ecosystems, riparian zones have been identified as important in controlling the export of sediment bound-P in surface runoff (Croke et al. 1999; Prosser et al. 2001). There is also evidence that riparian zones play a similar role for controlling the export of groundwater-borne nutrients to surface water (Lawrence et al. 1984). Much of the research on groundwater in the riparian zone has focused on the attenuation of NO3

–, the most common agricultural contaminant of groundwater (Lawrence 1983; Bolger and Stevens 1999). However, almost all the research on nutrient cycles in riparian groundwater has been made in humid temperate climates (Hill 1996, 2000; Burt et al. 2002). The role of the riparian zone in the attenuation in groundwater-borne contaminants under Australian conditions is poorly understood. This report is a part of a multidisciplinary project on groundwater–surface water interactions in a subtropical alluvial aquifer system (Wollombi Brook, NSW). The goal of the study was to provide some basic understanding on how groundwater–surface water interactions can impact on the load of nutrients to surface water in subtropical catchments. Wollombi Brook is typical of many southeastern Australian river systems, with an extremely variable flow regime and the occurrence of erosional sand deposits (“sand slugs”) in the river channel. Catchment-scale groundwater–surface water interactions and the regional hydrogeology of the Wollombi Brook have been summarised in a companion report (Herczeg et al. 2003). This report evaluates the significance of groundwater–surface water interactions on nutrient cycles in the brook. The evaluation of groundwater–surface water interactions and nutrient concentrations at different scales in a stream–alluvial aquifer system was challenging because of large spatial variability and complex interactions. Therefore, a substantial effort was spent developing appropriate equipment and methods during the study. These included an

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adaptation of the point dilution test to shallow alluvial aquifers (Lamontagne et al. 2002a) and several piezometer designs and porewater profilers (Appendix 3). A trial in-stream tracer addition was made on a tributary of the Wollombi Brook (Parsons Creek) to test the applicability of this technique in Australian streams (Appendix 4). Literature Review Riparian and hyporheic zones The terminology surrounding the concept of “riparian zone” is quite confused in the literature. Riparian zones can be defined as the region bordering the active channel that support long-lived vegetation (Dent et al. 2000). The riparian zone can be separated from surface water by the parafluvial zone (also called alluvial plain, fluvial plain, etc), which is the region of the active channel without surface water at low discharge (Dent et al. 2000; Woessner 2000). While “riparian zone” is often used synonymously with “floodplain” in the literature (Burt et al. 2002), it may be preferable to identify riparian zones as a floodplain component. In parallel to research on riparian zones, a large knowledge base is developing on the role of the hyporheic zone for nutrient processing in streams (Boulton et al. 1998; Boulton 2000; Jones and Mulholland 2000; Hinkle et al. 2001). The hyporheic zone can be defined as the region of saturated sediment where surface water and groundwater is actively mixing and exchanged (Boulton et al. 1998; Dent et al. 2000). Hyporheic processes occur at a variety of scales, from the small scale exchanges caused by obstacles along the stream bottom (Hutchinson and Webster 1998) to the transit of surface water through buried paleochannels (Huggenberger et al. 1998; Woessner 2000). While they form a continuum in scale of groundwater–surface water interaction, “riparian” and “hyporheic” processes are often studied separately. The integration of the different scales of groundwater–surface water interactions in hydrological and biogeochemical models is an active area of research (Jones and Mulholland 2000). Functional types of riparian zones Under temperate climates, some consensus is emerging on the functional types of riparian zones (i.e. those that can and cannot significantly impact nutrient concentration during the transit of groundwater). The impact of the riparian zone on nutrient biogeochemistry appears greatest when groundwater flow occurs primarily through shallow, organic-rich, soil horizons (Hill 1996, 2000; Hill et al. 2000). This form of riparian zone is more biogeochemically active because of increased potential for plant uptake and higher microbial activity. Burt et al. (2002) have hypothesised that such conditions for groundwater flow are most often encountered in U-shaped riverine corridors (i.e. those with a well-developed floodplain) as opposed to V-shaped ones or gullies (Fig. 1). These considerations about the shape of the riparian zone are currently included in models seeking to predict nutrient retention by riparian zones at the catchment-scale (Baker et al. 2001; Gold et al. 2001). While the original emphasis of research was on NO3

– attenuation, it is now recognized that riparian zones are primarily sites for nutrient transformations, that is, some forms of a nutrient will be lost and others will be gained during transit (Lamontagne et al. 2001; Burt et al. 2002). In practical terms, this means that an efficient removal of NO3

– by a riparian zone, for example, could be partially offset by a net export of other N forms (such as ammonium and dissolved organic nitrogen).

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The role of the riparian zone on catchment nutrient cycles may be more complex than originally thought. Many models of biogeochemical transformation during groundwater transit through the riparian zone assume that flow is principally from the hillslope to the river (Hill 1996). However, in some riparian zones, the main exchange between groundwater and surface water is through cycles of bank recharge and discharge during floods (Jolly et al. 1994; Martí et al. 2000; Rassam et al. 2002). Bank recharge–discharge cycles appear especially important in catchments with well-developed floodplains (Burt et al. 2002) or that are prone to flash floods (Martí et al. 2000). Most catchments probably exhibit (in space or time) some features of these two “models” of groundwater–surface water interaction in the riparian zone (Burt et al. 2002).

Figure 1: Functional types of riparian zones. a) U-shaped valley with a well-developed floodplain; b) V-shape valley; c) gully. Adapted from Burt et al. (2002).

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Figure 2: Location of Wollombi Brook (NSW) and the main sampling sites.

Methods Site description The Wollombi Brook is part of the Hunter River catchment of east Central NSW (Fig. 2). It has a sub-tropical climate (with predominantly summer rainfall) ranging from 1110 mm at Coorabong in the upper reaches of the catchment to 650 mm at Broke. The brook is the last major tributary of the Hunter River that has not been regulated. Much of the upper part of the catchment is pristine or plantation forest, with steep rocky slopes. Activities in the valleys include coal mining, cattle grazing, and irrigated agriculture (grapes, olives, and pastures). Triassic sandstone is the most common rock type in the upper part of the catchment (Erskine 1996). A variety of other rock types (including Quaternary alluvium, basalt, conglomerate, and coal) also occur within the lower reaches. In the valleys, the brook is associated with an alluvial aquifer. Because the regional groundwater is generally brackish (800 – 8 000 µS cm–1), the river and the alluvial aquifer are important supplies of freshwater for irrigation, stock watering, and private use. The hydrograph of the Wollombi Brook catchment is highly variable, with prolonged low flow periods and infrequent, short-lived, but potentially catastrophic floods (Erskine 1996). River morphology has changed substantially since European settlement, as described in detail by Scott and Erskine (1994) and Erskine (1996). Simply put, there has been substantial erosion of riverbanks and deposition of sand slugs in the river channel, thereby in-filling the river valley. Currently, the river channel and the floodplain are still recovering from the last major flood in 1949 (Erskine 1996). Hydrology In the first part of the study, the contribution of three potential sources of water to streamflow in Wollombi Brook was investigated (Herczeg et al. 2003). Potential sources included regional groundwater, alluvial groundwater, and delayed drainage of surface water. Using a variety of techniques, it was found that regional groundwater was a source of baseflow in the upper reaches but that the alluvial aquifer was a more important source of baseflow in the lower reaches. While a small flux of water, the discharge of saline regional groundwater in the lower reaches contributed to local increases in salinity in the deeper sections of the alluvial aquifer. Dewatering of coal seams during mining operations may have reduced or reversed the discharge of regional groundwater in parts of the lower reaches (Umwelt (Australia) Pty Ltd 2001).

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The water table in the alluvial aquifer is essentially a reflection of river level, with very weak gradients towards or away from the brook at baseflow. During storms, water level in the alluvial aquifer adjusts within hours to changes in river level. Bank recharge during floods can generate groundwater mounds where the alluvial plain meets the riparian slope. The dissipation of the groundwater mounds following floods appears an important component of the baseflow. The salinity of groundwater stored in the alluvial aquifer tended to increase during inter-flood periods, suggesting that evaporation from shallow water tables may occur. Riparian transects Three sites were selected along the bottom half of the catchment to study the exchange of water and nutrients between the regional aquifer, the alluvial aquifer, and the stream. The site near the town of Wollombi is immediately downstream from the confluence of two major tributaries and is at the transition between the upper and lower section of the brook (Fig. 3). The Fordwich and Warkworth sites are within the lower section of the brook and are characterized by lower gradients, wider alluvial plains, and large accumulations of sediments. The sites have common features including raised benched areas (or terraces), a sparsely vegetated alluvial plain, and a meandering or braided (Fordwich) river channel (Fig. 4). In January to May 2000, each site was instrumented with two deep boreholes installed at the margin of the upper terrace and with a network of nested piezometers in the alluvial plain. Bores were installed using a rotary drill. At each site, one shallow bore (15 to 17 m below the surface) was located at the interface between the upper unconfined and the regional bedrock aquifer. The second bore (50 m below the surface) tapped the deeper regional aquifer. Additional details about bore construction can be found in Herczeg et al. (2003). Piezometers were installed at five to six locations across the alluvial plain, usually in nests of two or three. The piezometers were made of 5.0 cm ID PVC slotted along the

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Figure 3: Cross-section of the Wollombi, Fordwich, and Warkworth sites. Labels with "DP" are drive points and those with "BP" are mini-piezometer bundles.

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lower 50 cm and covered with an outer lining to restrict the input of sand and silt during water sampling. Because of the non-cohesive nature of the alluvial sands, it was logistically difficult to install piezometers at depths greater than four metres. Several piezometers were lost during the course of the study, mostly because of damage during floods. In October 2000, the piezometer network was supplemented at each site with two to three drive points at depths ranging from three to six metres (i.e., usually near the bottom of the alluvial aquifer). In addition, mini-piezometer bundles were tested at the Wollombi and Fordwich sites in October 2000 and three were installed permanently at the Warkworth site in May 2001. Mini-piezometer bundles were designed to collect groundwater samples at 25 to 50 cm intervals for up to three metres below the water table. More details on the construction, installation, and sampling of drive points and mini-piezometers are given in Appendix 3. The thickness of the alluvial aquifer along the transects was measured using steel rods. It was assumed that the depth to hard contact represented either bedrock or the former river channel. The depths of contact obtained with the steel rods in the alluvial aquifer were similar to the depth of the bedrock recorded during the drilling of the bores (Fig. 3 and Herczeg et al. 2003). Sampling Techniques Bores and piezometers On the first days of each sampling trip, the water level was measured and the bores and piezometers were pumped to dryness or were flushed for several well volumes. A copious amount of Fe oxide precipitates was often found in the piezometers during early flushing stages. On the following day, groundwater was collected with minimal exposure to the atmosphere by inserting the end of the pump line into the bottom of a large Erlenmeyer flask and by letting it overflow. Field EC, pH, dissolved oxygen, and temperature were measured using a WTW Multiline P4 Universal meter. Sub-samples were collected from the bottom of the flask using a large syringe connected to a three-way valve. Samples were immediately filtered on-line with a 0.45 µm Supor© membrane filter (Pall). A small filtered subsample was collected for Cl– analysis in a scintillation vial. Another subsample was added to a small pre-prepared vial containing the reagents for Fe2+ analysis. The remaining filtered water was collected in a 125-mL bottle and acidified to pH<2 using analytical grade HCl for major ions and nutrients (Table 1). Alkalinity was usually measured on site using a field titration kit (Hach). Prior to each sampling trip, all sampling bottles and filtering equipment were washed in P-free detergent and in a mild acid bath before thorough rinsing with distilled deionised water. For piezometers that were slow to recover, it was not always possible to remove three piezometer volumes

Figure 4: Plane view of the Warkworth site. The A – A’ line represents the approximate location of the cross-sectional transect shown on Fig. 4.

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Table 1: Major ions and nutrients measured in water quality samples collected from river and groundwater. Not all elements were measured for each sampling trip or for each piezometer during some trips. Analytical methods are outlined in Appendix 1.

Analyte

Acronym

Ca2+, Mg2+, Na+, K+, Cl–, SO42-

Alkalinity

Fe2+ Filterable Reactive P (“phosphate”) FRP Total Dissolved Phosphorus TDP Total Dissolved Nitrogen TDN Ammonium (NH4

+) Nitrate (NO3

–) Dissolved Organic Carbon DOC Dissolved Organic Nitrogen DON Filterable Reactive silica FRSi Total Dissolved Iron TDFe

prior to sample collection and sampling commenced soon after pumping on the second day. Analytical methods are outlined in Appendix 1. Drive points Drive points were pumped dry on the day prior to sampling using a sipper (6 mm OD nylon tubing) and a hand pump. Samples were collected in a 125-mL bottle on the following day and handled as described above. Mini-piezometers Individual mini-piezometers were pumped for several volumes on the day prior to sampling. Samples were collected on the following day and handled as described above. It was subsequently discovered that the nylon tubing used to build the mini-piezometers contaminated groundwater samples with DOC and DON (Appendix 3). DOC and DON results from the mini-piezomters and the drive points (which were also sampled with a length of the nylon tubing) will not be further discussed here. Stream porewater profiles During the March–April 2001 and November 2001 field trips, porewater profiles were taken at the Warkworth site using the stream profiler (Appendix 3). At each sampling date, a transect was established across the brook and water depth measured every 1 m. Stream velocity was measured every one or two metres using a current meter or (at very low flows) estimated using the time required by a floating object to travel a certain distance downstream. Duplicate porewater profiles were measured at mid-stream and along one of the shorelines. For each profile, samples were collected every 10 cm from the surface of the streambed to one meter depth. Samples for selected analytes were handled and processed as described above. Following the collection of the porewater profiles, hydraulic head was measured at selected locations with a potentiomanometer (Winter et al. 1988). The potentiomanometer compares the difference in hydraulic head between the stream and a point in the subsurface (at a depth of 1 m in this case). Comparison of alluvial groundwater discharge and hyporheic exchange to the nutrient budget of Wollombi Brook The load of nutrients from groundwater at Wollombi Brook will be some combination of the load from alluvial groundwater discharge and the one from hyporheic exchange. The relative importance of both processes was compared using the stream flow, hydraulic head, and porewater profiles collected at the stream cross-section at Warkworth in April 2000. The volume of groundwater discharge along the stream cross-section (assuming a 1-m width of section) was estimated using:

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AiKQ BG= (1),

where QG is groundwater discharge (m3 s–1), K the hydraulic conductivity (m s–1), i the hydraulic gradient, and AB the surface area of the streambed (m2). K was estimated using a point dilution test (Lamontagne et al. 2002a) at varying between 0.5 – 1x10–3 m s–1 (consistent with a fine sand porous medium; Freeze and Cherry 1979). Because of differences in hydraulic head along the stream cross-section, QG was first estimated for each 1–m interval (i was extrapolated from nearby intervals when not measured) and then summed for the cross-section. A range of possible volume of hyporheic exchange was derived from general relationships for sandy streambeds (Harvey and Wagner 2000). The hyporheic flux can be defined as:

Aqs α= (2),

where qs is the storage exchange flux from water to sediment per unit stream length (m3 s–1 m–1), α the storage exchange coefficient (s–1), and A the cross-sectional area of the stream (Harvey and Wagner 2000). Because hyporheic exchange involves no net flux of water in or out of the streambed, qs will be equal in magnitude but opposite in sign to qh, the hyporheic-exchange flux from sediment to water. The exchange coefficient can be determined using:

A

At SS α

= (3),

where ts is the average residence time in the hyporheic zone and As the stream storage area (assumed here to be only the hyporheic zone). The range in ts for sandbed forms was estimated at 100 to 2000 s–1 (Harvey and Wagner 2000). As was estimated from the apparent depth of mixing between streamwater and groundwater in the porewater profiles. The hyporheic-exchange flux can be estimated by substituting equation 3 in equation 2:

tAq

S

SS

= (4).

Nitrogen mineralisation To assess the location for potential sources of N to groundwater across the floodplain, rates of nitrogen mineralisation were estimated at the Wollombi transect on one occasion (October 2000) using the buried bag method (Eno 1960). Briefly, this method estimates the rates of NH4

+ and NO3– production in soil samples incubated in situ in polyethylene

bags. Bags were incubated in triplicate at several depths at four stations spanning the riparian zone. For stations on alluvial sand, bags were incubated just below the soil surface, above, and below the water table (except for the station closest to the brook

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Figure 5: Rainfall (bar graph) at Broke and stream level flucuations at Warkworth (dots) from Jan 2000 to Feb 2001. Timing of field trips indicated by dashed lines.

where bags could only be incubated above and below the water table). At a site on the terrace, the LF horizon and the underlying sandy C horizon were incubated. Buried bags were incubated for 10 days. Ammonium and NO3

- were extracted from soils subsamples using 2 M KCl (Lamontagne et al. 2002b). Net N mineralisation rates (in µg N (g soil)–1 day–1) and net nitrification were estimated from the changes in mineral N and NO3

– concentration between final and initial bags, respectively (Lamontagne 1998). Field Trips The period of study spanned from January 2000 to February 2001. Seven sampling trips were made during this period, covering different flow conditions (Fig. 5). However, no

sampling period coincided with peak flow conditions. In addition to the standard monitoring of piezometers, run-of-river surveys for environmental tracers were made on three occasions and a bi-weekly monitoring program was made at one station from May 2000 to February 2001. These are outlined in more detail in the companion report (Herczeg et al. 2003). Detailed sampling of the piezometer network for water quality parametres was not done for every field trip (see Appendix 2). Results General patterns in water quality A wide variety of geochemical environments were present across the alluvial plain environment but some general trends emerged between regional groundwater (as represented by the bores), alluvial groundwater, and river water (Table 2). Groundwater from the regional aquifer and the alluvial aquifer tended to be sub-oxic to anoxic, whereas river water was well oxygenated (Table 2). A wide range in salinity (as represented by EC or Cl– concentration) was found in all three types of water. However, regional groundwater was usually brackish (736 – 7570 µS cm–1) whereas alluvial

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 10

groundwater (155 – 6430 µS cm–1) and river water (388 – 1377 µS cm–1) ranged from fresh to occasionally brackish. In all water types, pH was circumneutral to alkaline (6.1 to 8.0). Some extremely high alkalinities (>10 meq L–1) and Na+ concentrations (>1000 mg L–1) in one of the regional bores (79059) may have been caused, in part, by contamination from the bentonite seal used to isolate the gravel pack from downward leakage. Nutrient concentrations (N, P, and Si) were higher in groundwater than in surface water. For nitrogen, both mineral (NH4

+ and NO3–) and organic forms (DON) were important

(Table 2). NH4+ and DON were the main forms of dissolved N in groundwater while DON

was the most common form in surface water. For phosphorus, FRP could account for most of the dissolved P in groundwater. In surface water, FRP was always below the detection limit and most of the P was in the dissolved organic fraction (i.e. TDP – FRP ~ Dissolved Organic P). The range in DOC concentration in alluvial groundwater and in surface water was similar and higher than in regional groundwater (Table 2).

Spatial trends across the alluvial plain A large database was accumulated on water quality trends along the three alluvial plain transects (Appendix 2). For the sake of brevity, an emphasis will be given to the March–

Table 2: Mean and range in water quality parameters from regional groundwater, alluvial groundwater, and river water for all sampling trips. The true range may be larger for some water types (stream water) because not all flow conditions were sampled during the study. Data in mg L-1 unless otherwise noted. Data from the Warkworth mini-piezometers not included to avoid a bias towards conditions at that site. Bores

(Regional groundwater)

Alluvial groundwater River

Temp (°C) 19.6 (9.3 – 23.8) 19.6 (9.4 – 25.3) 13.6 (8.5 – 22) Field pH 6.9 (6.1 – 7.6) 6.8 (6.0 – 7.7) 7.4 (6.4 – 8.0) Dissolved oxygen 0.2 (<0.1 – 0.4) 0.4 (<0.1 – 2.4) 8.3 (7.4 – 9.2) EC (µS cm–1) 2259 (736 – 7570) 714 (155 – 6430) 597 (388 – 1377) Alkalinity (meq L–1) 8.8 (0.64 – 19) 2.0 (0.35 – 8.8) 1.8 (0.97 – 3.2) Ca2+ 38 (11 – 75) 17 (2.3 – 67) 16 (9.9 – 29) K+ 8.8 (0.9 – 26) 4.0 (1.6 – 13) 5.6 (3.9 – 9.2) Mg2+ 28 (8.2 – 67) 14 (2.5 – 73) 21 (10 – 44) Na+ 521 (90 – 2100) 79 (17 – 649) 126 (49 – 320) Cl– 308 (71 – 1285) 141 (13 – 1740) 194 (86 – 440) SO4

2– 17 (<0.2 – 42) 14 (<0.2 – 136) 24 (7.3 – 88) NH4

+ (µg N L–1) 860 (<20 – 2200) 460 (<20 – 2170) 67 (<20 – 280) NO3

- (µg N L–1) 160 (<5 – 1700) 140 (<5 – 4600) 45 (<5 – 190) DON (µg N L–1) 260 (<50 – 520) 310 (110 – 530) 380 (190 – 570) TDP (µg L–1) 129 (17 – 320) 56 (<5 – 480) 12 (<5 – 16) FRP (µg L–1) 98 (17 – 260) 51 (<5 – 470) <5 Fe2+ 8.5 (<0.01 – 40) 11 (<0.01 – 44) 0.20 (<0.01 – 0.30) TDFe 6.9 (0.02 – 54) 11 (<0.1 – 126?) 0.27 (0.040 – 0.89) DOC 2.3 (<0.5 – 6.7) 4.5 (2.5 – 14) 6.3 (4.9 – 8.4) FRSi 10 (6.0 – 25) 5.4 (1.6 – 14) 2.1 (2.1 – 3.0)

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 11

April 2001 trip (which was at the end of flood recession following a storm) and the November 2001 trip (which occurred after a prolonged period of baseflow). Redox environment The most important factor determining the speciation and concentration of nutrients in groundwater was the redox potential of the aquifer, especially the presence of oxygen. However, oxygen measurements using the probe were not accurate below 1 mg L–1 and the probe itself frequently broke down or required calibration (possibly because of the presence of strong reductant in many samples). In addition, the probe was not practical when only a small amount of water could be extracted (i.e. mini-piezometers). The measurement of Fe2+ was selected as an alternative to O2 to assess the redox potential of groundwater samples because only a few milliliters were required for the analysis. The presence of Fe2+ in water samples indicates that water is sub-oxic to anoxic. TDFe measurements were used as a substitute to Fe2+ when the latter was not available. At the range in pH found in groundwater at the sites, dissolved Fe will be mostly as Fe2+ (Appelo and Postma 1993). In general, alluvial groundwater at the Wollombi (Fig. 6) and Fordwich sites (Fig. 7) was strongly reduced, while oxygenated pockets (areas with Fe2+ below detection limit) were present at Warkworth (Fig. 8). Why Warkworth is different than the other sites is not clear but may be related to a deeper water table in the alluvial plain. During the March– April 2001 sampling period, for example, the water table was at the surface at Wollombi and at Fordwich. In contrast, depth to the water table ranged between one or two metres below the surface at Warkworth. At Warkworth, Fe2+ concentrations were lowest near the tail of the slope and highest closer to the river. Unexpectedly, small concentrations of Fe2+ were also occasionally found in river water. Fe2+ in river water may be derived from hyporheic exchange with stream sediments or from the discharge of alluvial groundwater (see Porewater profiles below).

––

2

30*–30

39

27

20

14

20

a) March – April 2001

∇ ∇

––

31*–23*

27*

44

15*

2

40

∇ ∇

b) November 2001

0.2

––

2

30*–30

39

27

20

14

20

a) March – April 2001

∇ ∇

––

31*–23*

27*

44

15*

2

40

∇ ∇

b) November 2001

0.2

Figure 6: Fe2+ concentration (mg L–1) at Wollombi Brook (data with ‘*’ estimated from TDFe concentration).

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 12

3

3

0.911

19

3

0.6

22–

<0.1

8

5

213

19

6

6

42<0.01

0.3

∇ ∇

0.1

∇ ∇

b) November 2001

a) March – April 2001

3

3

0.911

19

3

0.6

22–

<0.1

8

5

213

19

6

6

42<0.01

0.3

∇ ∇

0.1

∇ ∇

b) November 2001

a) March – April 2001

Figure 7: Fe2+ concentration (mg L–1) at Fordwich.

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 13

Patterns in nutrient concentration The patterns in nutrient concentration in groundwater were consistent with the redox conditions inferred from Fe2+. Ammonium and NO3

– concentration had opposite patterns as a function of Fe2+. High NO3

– concentrations were only found when Fe2+ was low or below detection limit, whereas NH4

+ concentration had a weak tendency to be higher at higher Fe2+ concentrations (Fig. 9). There were also some distinct spatial and temporal patterns in NH4

+ and NO3– concentrations at some sites.

At Warkworth, a pocket of oxygenated groundwater containing NO3

– occurred at the toe of the slope (Fig. 10). The extent of this pocket tended to vary from one sampling

∇∇

0.15

0.1

<0.1

16

6

11

9

6

0.2

<0.1

–<0.1<0.1<0.1<0.1<0.1

<0.1<0.1<0.1

––

<0.1

<0.1<0.1<0.1<0.1<0.1

0.2

a) March – April 2001

∇∇

0.54

<0.1<0.1

0.10.9

8

17

6

14

<0.1

<0.1

<0.1<0.1

3.60.2

<0.1

––

<0.1<0.1

<0.1<0.1<0.1<0.1

b) November 2001

∇∇

0.15

0.1

<0.1

16

6

11

9

6

0.2

<0.1

–<0.1<0.1<0.1<0.1<0.1

<0.1<0.1<0.1

––

<0.1

<0.1<0.1<0.1<0.1<0.1

0.2

a) March – April 2001

∇∇

0.54

<0.1<0.1

0.10.9

8

17

6

14

<0.1

<0.1

<0.1<0.1

3.60.2

<0.1

––

<0.1<0.1

<0.1<0.1<0.1<0.1

b) November 2001

Figure 8: Fe2+ concentration (mg L–1) at Warkworth. Arrows represent groundwater flowpaths.

Figure 9: Ammonium and nitrate concentrations as a function of Fe2+ in alluvial groundwater.

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 14

period to another (Fig. 10) but it was always present. In contrast, groundwater closer to the river was always sub-oxic to anoxic and only contained trace amounts of NO3

–. Because of the analytical method used (which measures NO3

– + NO2–) trace ‘NO3

–‘ concentrations may be mostly NO2

–. Elevated NO3– concentrations were never found in

piezometers close to the brook at all transects but elevated NH4+ concentrations (>0.5

mg N L–1) were common (Fig. 11; Appendix 2).

∇∇

10010

210

100

––

8

70

40

30

40

160

–907

<5406

–140240––

1500

110560052002200

250220

a) March – April 2001

∇∇

4020

4600<5

<510

190

80

10

20

<5

470

11003800

70570440

––

46004600

2100370040004600

b) November 2001

∇∇

10010

210

100

––

8

70

40

30

40

160

–907

<5406

–140240––

1500

110560052002200

250220

a) March – April 2001

∇∇

4020

4600<5

<510

190

80

10

20

<5

470

11003800

70570440

––

46004600

2100370040004600

b) November 2001

Figure 10: Nitrate concentration (in µg N L–1) at the Warkworth transect on a)March–April 2001 and b) November 2001.

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 15

Unlike the nitrogen species, the distribution of FRP was not well correlated with Fe2+ (Fig. 12). FRP concentration was usually low when Fe2+ was less than 1 mg L–1. However, both high and low FRP concentrations were found in samples with high Fe2+. At Warkworth, FRP concentration tended to increase following a prolonged period of baseflow (Fig. 13). On average, FRP concentrations in alluvial groundwater were generally higher than in the stream water (which always had FRP <5 µg L–1).

∇∇

190440

210

190

––

980

240

980

540

370

450

–20

<20206020

–40

<20––20

50<20<201203020

a) March – April 2001

∇∇

120150

7060

200350

500

280

540

220

270

1200

12040

1306050

––2030

2020

<2020

b) November 2001

∇∇

190440

210

190

––

980

240

980

540

370

450

–20

<20206020

–40

<20––20

50<20<201203020

a) March – April 2001

∇∇

120150

7060

200350

500

280

540

220

270

1200

12040

1306050

––2030

2020

<2020

b) November 2001

Figure 11: Ammonium (in µg N L–1) at the Warkworth transect in a) March–April 2001 and b) November 2001.

Figure 12: Relationship between Fe2+ and FRP in alluvial groundwater.

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 16

∇∇

<523

<5

<5

––

<5

<5

9

<5

14

12

–<5<5<5<5<5

–<5<5––

<5

<5<5<5<5<5<5

a) March – April 2001

∇∇

1224

21<5

5<5

12

<5

15

10

24

36

57

<577

––66

<5<5<510

b) November 2001

∇∇

<523

<5

<5

––

<5

<5

9

<5

14

12

–<5<5<5<5<5

–<5<5––

<5

<5<5<5<5<5<5

a) March – April 2001

∇∇

1224

21<5

5<5

12

<5

15

10

24

36

57

<577

––66

<5<5<510

b) November 2001

Figure 13: FRP concentration (µg L-1) at Warkworth on a) March–April 2001 and b)November 2001.

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 17

Water Table profiles Even when the alluvial aquifer was primarily anoxic, it was suspected that an oxygenated fringe could occur at the interface between the unsaturated zone and the water table. This hypothesis was tested at the Wollombi and Fordwich transects with nests of minipiezometers designed to collect groundwater profiles at close intervals near the water table. An oxygenated fringe (here defined by a sharp gradient in TDFe concentration) was present at three of the four profiles (Fig. 14). When present, the oxygenated fringe extended to 20 to 40 cm below the water table. Gradients in NH4

+ and occasionally FRP concentration were also found when an oxygenated fringe was present (Fig. 14).

Figure 14: Water table profiles for selected chemical species taken in October 2000 at the Wollombi and Fordwich transect.

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 18

Streambed porewater profiles The biogeochemical transformations in the hyporheic zone were investigated with porewater profile sampling at Warkworth during the March–April 2001 and November 2001 trips. On the first trip, two profiles were taken at the stream edge and one at mid-stream. On the second trip, two profiles were taken at the stream edge and two at mid-stream. In addition to water samples, hydraulic gradients in the vicinity of each profile were measured using a potentiomanometer (Winter et al. 1988) and stream velocity measured using a current meter. Potentiomanometry was also used to measure hydraulic gradients between piezometers and the stream. The nature of the groundwater–surface water interaction was different between the two sampling dates. In April 2001 (following the flood), stream velocity was relatively high (~0.4 m s–1) and the hydraulic gradient in the alluvial plain was toward the stream (0.006 m/m). Hydraulic gradients indicated groundwater discharge throughout the streambed, with larger gradients at the right bank (Table 3). In contrast, in November 2001 stream velocity was low and a smaller hydraulic gradient from the alluvial plain to the stream was present (0.0005 m/m). However, in the streambed, hydraulic gradients were away from the stream (i.e. the stream was “losing” water; Table 3). Table 3: Results of stream cross-section surveys at Warkworth in April 2001 and November 2001. All distances measured from the left bank. The arrows on the hydraulic gradients indicate the direction of groundwater flow.

5 April 2001

7 November 2001

Distance (m from left bk)

Stream Depth (cm)

Velocity (m s–1)

Hydraulicgradient

Distance Stream Depth (cm)

Velocity (m s–1)

Hydraulic gradient

0 0 0 Mid-river <20 <0.01 0.0033 ↓ 1 20 0.069 0 (~6 m) <20 <0.01 0.0053 ↓ 2 49 0.16 3 50 0.28 0.0019 ↑ Right bk. <1 0 0.0025 ↓ 4 49 0.38 (~12 m) <1 0 0.0020 ↓ 5 44 6 40 0.37 7 37 0.0013 ↑ 8 28 0.28 9 28

10 26 0.23 11 24 0.0030 ↑ 12 23 0.22 13 20 14 15 0.12 0.0031 ↑ 15 <1 0 0.0042 ↑

These differences in hydrological exchange were associated with significant contrasts in the porewater profiles. In April 2001, sediments were well oxygenated down to 40-cm depth (based on Fe2+ concentration) at the mid-stream profile but anoxic throughout the shoreline profiles (Fig. 15). This suggests a well-developed hyporheic zone at mid-stream (where groundwater mixes with surface water before reaching the surface) and direct discharge of alluvial groundwater along the shoreline. The patterns in groundwater

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 19

and surface water mixing were also reflected in some other chemical species. For example, the porewater NH4

+ profile at the shoreline was consistent with groundwater input with little hyporheic mixing, whereas at the mid-stream location NH4

+ concentration varied from “streamwater”-like at 10 cm depth to “groundwater”-like at 60 cm-depth (Fig. 15). The patterns in conservative tracers (such as Cl–) were also consistent with different origins for the porewater found at mid-stream and shoreline locations. In November 2001, the stream appeared to be a groundwater flow-through system instead of a groundwater discharge system. While a weak gradient from the alluvial plain to the stream was present, potentiomanometric measurements along the streambed suggested that the stream was recharging the alluvial aquifer. The patterns in Fe2+ and Cl– were markedly different between the shoreline and mid-stream profiles (Fig. 16), also suggesting that not all porewaters originated from river recharge. Overall, the patterns in potentiomanometry and porewater profiles suggest that a weak input of low Cl–, oxygenated groundwater was occurring at the very edge of the stream (possibly bank discharge of streamwater from a small, recent spate). However, most of the streambed appeared to be losing water. The increase in Cl– at the base of the shoreline porewater profiles suggests that a mixing zone between discharging alluvial groundwater and recharging streamwater occurs at about 60 cm depth (Figs. 16 and 17).

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 20

Figure 15: Porewater profiles for selected chemical species at Warworth, 5 April 2001. Arrows on the hydraulic gradients indicate the direction of groundwater flow.

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 21

Figure 16: Porewater profiles for selected chemical species at Warkworth on 6 November 2001.

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 22

Figure 17: Conceptual model of stream-groundwater interaction at the Warkworth transect on a) March–April 2001 (a groundwater discharge system with hyporheic mixing) and b) November 2001 (a local groundwater discharge–regional rechargesystem).

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 23

Comparison of potential water fluxes from alluvial groundwater discharge and hyporheic exchange The comparison of alluvial groundwater discharge and hyporheic exchange flux suggests that hyporheic exchange dominated groundwater–surface water interactions at Warkworth on 5 April 2001 (Table 4). The hyporheic flux estimates required the measurement of the storage zone cross-sectional area (As – assumed to be only the hyporheic zone here). Based on the porewater profiles, the depth of active hyporheic mixing (dh) was less than 10 cm near the banks but possibly as deep as 50 cm at mid-stream (Fig. 15). It was estimated that, on average for the stream cross-section, dh ranged between 10 to 30 cm. Thus, assuming no bank exchange and for a 15 m stream width, As ranged between 1.5 to 4.5 m2. The estimated hyporheic-exchange flux ranged between 7.5 x 10–4 – 4.5 x 10–2 m3 s-1 per meter of stream length (Table 4). In contrast, alluvial groundwater discharge per meter of stream length ranged between 1.5 – 3.0 x 10-5 m3 s–1 (Table 4). Thus, the hyporheic-exchange flux was one to three orders of magnitude larger than the alluvial groundwater flux.

Table 4: Comparison of alluvial groundwater flux and hyporheic-exchange flux at Warkworth on 5 April 2001 for a range in parameter values.

Flux

(m3 s–1 per m of stream length) Alluvial groundwater K = 5x10–4 m s–1 1.5 x 10–5 K = 1x10–3 m s–1 3.0 x 10–5 Hyporheic exchange As = 1.5 m2, ts = 2000 s–1 7.5 x 10–4 As = 4.5 m2, ts = 100 s–1 4.5 x 10–2

The stream length required for one volume of surface water to be exchanged with the hyporheic zone (or turnover length – Ls) is a good way to visualise the possible significance of hyporheic exchange to nutrient fluxes (Harvey and Wagner 2000). The turnover length can be estimated using:

αuLS = (5),

where u is the average stream velocity (i.e. the volumetric flow rate divided by the flow cross-sectional area). Estimates for Ls for the range in α from the Warkworth cross-section (1.8 x 10–4 to 1.1 x 10–2 s–1) vary between 25 to 1500 m. This implies that stream water will be exchanged many times with the streambed by hyporheic exchange during its transit through the catchment. A similar measure for alluvial groundwater discharge (the length of stream required to double the volumetric flow rate) ranges from 40 to 80 km. This is consistent with findings from the first part of the study, which determined that most of the stream water in the lower reaches was derived from the alluvial aquifer during low flows (Herczeg et al. 2003).

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 24

The contribution of alluvial groundwater discharge to the nutrient budget of the brook may be more significant than volumetric discharge per se because some nutrients have a relatively high concentration in groundwater. This can be evaluated from the hypothetical stream length required to double the flux of nutrients in the brook from alluvial groundwater discharge (assuming no algal uptake, etc). The “doubling nutrient flux” distances are shorter than for groundwater itself. Using the average concentration for alluvial groundwater and streamwater (Table 2), “doubling flux” distances range from 1.5–3 km for Fe2+, 4–8 km for FRP, 7–14 km for NH4

+, and 16 – 32 km for SrSi. The fluxes of nutrients from alluvial groundwater defined here should be considered “potential fluxes” to the stream because several of the nutrients in groundwater could be chemically or biologically reactive in the hyporheic zone. At one extreme, assuming no reaction whatsoever, hyporheic exchange would only dilute nutrient concentrations by mixing streamwater with alluvial groundwater but would not alter the fluxes to the brook. At the other extreme, biogeochemical reaction in the hyporheic zone would impart a characteristic signature for nutrient concentrations regardless of the origin of hyporheic water (i.e. surface or groundwater). In the latter case, nutrient discharge to the stream will be the combination of groundwater discharge and hyporheic return times nutrient concentration in the hyporheic zone. The “actual” nutrient discharge from alluvial groundwater will be somewhere between these two scenarios. For example, Fe2+ in alluvial groundwater was apparently “lost” during hyporheic mixing at the mid-stream location (probably by oxidation to Fe3+ when exposed to O2). On the other hand, a significant Fe2+ input to the stream could still occur because a large proportion of the alluvial groundwater discharge occurred along the stream banks, where the hyporheic zone was thin or absent. It was not possible to quantify the net flux of nutrients from the streambed to the stream because the net result of hyporheic exchange (gain or loss of nutrients) could not be determined. Gain or loss of nutrient during hyporheic mixing could be deduced by comparing porewater nutrient profiles relative to the one of a conservative tracer (such as Cl– ; Officer 1979; Officer and Lynch 1981). However, surface and groundwater did not have distinct and constant end-members for Cl– in the profiles. However, the patterns in the speciation of redox-sensitive elements in stream water and pore water suggest that the net result of groundwater discharge and hyporheic exchange will be to favour the export of species more stable or soluble under reduced conditions (i.e. NH4

+ and Fe2+). Hyporheic exchange would also tend to remove some of the oxidized species found in stream water (i.e. NO3

–). This is consistent with the primarily reduced state of the alluvial aquifer in the vicinity of the brook.

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 25

Nitrogen mineralisation rates across the floodplain There was a distinct pattern in net N mineralisation and net nitrification rates across the floodplain at the Wollombi transect (Fig. 18). In general, rates were higher in the litter layer and shallow alluvial sands than in alluvial sand soil above or below the water table. Rates were lowest near the stream bank and highest at the toe of the slope and on the terrace. These patterns in N mineralisation suggest that a significant pool of soil NH4

+ and NO3

– could accumulate in shallow soil and alluvial sands between flooding or rainfall events. Thus, some of the NH4

+ and NO3– found in alluvial groundwater could originate

from the riparian zone rather than from the export of dissolved N in groundwater from further away in the landscape.

Figure 18: Net N mineralisation (MINnet) and net nitrification rates (NO3

-net)

at the Wollombi transect, October 2000 (in µg N g-1 d-1).

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Discussion In Wollombi Brook, groundwater–surface water interactions were driven by a combination of bank recharge and discharge cycles with the alluvial aquifer during floods and hyporheic exchange with the streambed at lower flows. Using several environmental tracers, Herczeg et al. (2003) determined that a large fraction of baseflow in the lower section of the brook originates from alluvial groundwater. While the lower section of the catchment is a regional groundwater discharge area, its contribution to baseflow is small. In addition, regional groundwater apparently seeps through the bottom of the alluvial aquifer (Herczeg et al. 2003), effectively by-passing the riparian zone. While the amount of nutrients delivered was probably small, saline regional groundwater discharge may still impact nutrient cycles in surface water by modifying the biogeochemical environment, for example through the generation of density stratification (Anderson and Morison 1989; Bailey and James 2000). The dominance of “lateral” exchange processes at Wollombi Brook has some of the attributes proposed in the “flood-pulse” theory for the interaction between rivers and their floodplains (Junk et al. 1989). According to the flood-pulse theory, the exchange of water, organic matter, and nutrients between rivers and floodplains occurs when surface water connectivity is re-established, principally during floods. However, unlike what is proposed in the flood-pulse theory, rivers and floodplains can remain connected following floods even when surface water connections are severed. The maintenance of a groundwater connection between floodplains and rivers can have important implications for salt and nutrient delivery to surface water. For example, in the Chowilla floodplain on the lower River Murray, an increased discharge of salt through groundwater persisted for 18 months following a medium-size flood (Jolly and Walker 1995). A similar process may occur for several nutrient species with elevated concentration in groundwater (especially FRP, NH4

+, SrSi, and Fe2+). The input of nutrients from groundwater may contribute to the pulse in biological activity observed in semi-arid rivers following floods but is yet to be accurately quantified. Nutrient recycling in groundwater during lateral exchange The cycling of nutrients during lateral exchange of groundwater and surface water is very active, in part because several redox interfaces can be encountered in space and time during the cycle from surface to groundwater and back (Bourg and Bertin 1993). For example, at least three redox interfaces were identified in the Wollombi Brook alluvial aquifer, including 1) where pockets of oxygenated groundwater moved into reduced sections of the alluvial aquifer, 2) at the water table oxycline, and 3) in the hyporheic zone. Thus, many opportunities exist for nutrients to be transformed (or lost) during transit through the alluvial aquifer. In other words, nutrients can be delivered back to surface water in a completely different form to that in which they were delivered to the floodplain (Lamontagne et al. 2001). This will be illustrated here using a sketch of the N cycle in the Wollombi Brook alluvial aquifer (Fig. 19). This sketch is drawn from our experience at Wollombi Brook, the River Murray floodplain (Lamontagne et al. 2002b) and other studies. Detailed summaries of the N cycle at the sediment–water interface (Baldwin and Mitchell 2000) and the hyporheic interface (Jones and Mulholland 2000) are available elsewhere. The emphasis here will be given to potential processes occurring in the unsaturated and saturated zone.

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Figure 19: Conceptual model of the nitrogen cycle during a lateral exchange cycle in the unsaturated and saturated zone of an alluvial aquifer.

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Sources of N to alluvial aquifers There are numerous potential sources of N to riparian zones, including atmospheric deposition, upland surface runoff, and regional groundwater discharge. However, in arid and semi-arid climates, sediment deposition during floods, nitrogen fixation, and floodwater recharge may be more significant. For some of these sources (groundwater discharge, floodwater recharge, etc) the input of dissolved N to the alluvial aquifer could occur directly. However, for other sources (N fixation, sediment deposition, etc), N will be mostly tied in biomass and in the soil organic matter pool. Nitrogen tied in soil and biomass will only become available to transport to alluvial groundwater as a by-product of the N cycle in the floodplain. There are two pathways by which N tied in particulate organic matter could enter the alluvial groundwater system. Part of the organic–N pool will be lost during recharge events by leaching of DON originating from the decomposition of litter and soil organic matter (Baldwin 1999; Robertson et al. 1999). The movement of DON through the unsaturated zone is not well understood but is thought to be primarily controlled by abiotic processes, such as adsorption on soil clay minerals (Kalbitz et al. 2000). The second internal pathway for dissolved N input to the alluvial aquifer will be through the mineralisation of organic–N to NH4

+ and NO3–. The

recycling of organic-N is a significant source of N for plant growth in riparian zones (Schade et al. 2002). However, when the rates of mineralisation exceed the demand by plants, a significant pool of mineral N can accumulate in soils (Aber et al. 1989). This excess mineral–N accumulating in the unsaturated zone could be displaced to groundwater during recharge events promoted by large rainfall or flooding. For example, at the Hattah floodplain on the River Murray, Lamontagne et al. (2002b) found that both Cl– and NO3

– accumulated in the unsaturated zone of the riparian zone between floods. However, while Cl– concentration increased in shallow groundwater following a large flood, NO3

– did not (consistent with a reducing environment in the aquifer at the site). Finally, in addition to N displaced from the unsaturated zone, part of the dissolved N pool found in alluvial groundwater probably originates from decomposition of organic matter buried in the aquifer. Mechanisms of NO3

– attenuation in alluvial aquifers While NO3

– was recharged to some sections of the Wollombi alluvial aquifer, practically all of it was apparently consumed before reaching the brook. Several biogeochemical pathways can be involved in the removal of NO3

– from groundwater. These include: 1) assimilatory reduction by plants, 2) assimilatory reduction by microorganisms, 3) dissimilatory reduction to NH4

+, and 4) dissimilatory reduction to N2 or N2O (denitrification). The knowledge of which processes are removing NO3

– from groundwater is critical because in some cases N is permanently removed (i.e. denitrification) while in others N is retained within the system but in another form. While all the above-mentioned processes can occur at least under some circumstances, it is, unfortunately, often assumed that NO3

- attenuation in riparian zones occurs principally through some combination of plant uptake and denitrification (Hill 1996). However, the relative importance of the two dissimilatory reduction processes is not well understood (Tobias et al. 1989). Denitrification is thought to be the principal process in freshwater or in environments with a variable aerobic regime (Tiedje et al. 1981; Højberg et al. 1994). In contrast, dissimilatory reduction of NO3

– to NH4+ appears more common

in salt marshes and in strongly anoxic marine sediments (Tiedje 1988). Because of the wide range in environmental conditions found in the Wollombi Brook aquifer (i.e. from fresh to saline), both processes could occur.

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Water table depth At both Hattah and Wollombi, a significant concentration of NO3

– in alluvial groundwater was only found where the water table was relatively deep. Similar patterns have been observed in sandy aquifers and have been associated with the movement of DOC in the unsaturated zone (Starr 1988; Pabich et al. 2001). Where water tables are deeper, most of the DOC produced in organic soil horizons is retained by sorption or consumed by microorganisms in the unsaturated zone before reaching the water table. As a result, not enough labile DOC is found in groundwater to produce the sub-oxic to anoxic conditions required for nitrate reduction. In contrast, where water tables are shallower, more DOC can reach the water table and will be available for microbial respiration. In addition to increased DOC inputs, shallow floodplain sediments can be enriched in organic deposits that can be used as a substrate for bacterial respiration (Hill et al. 2000). Similarly, Pinay et al. (2000) also found that the denitrification potential in a range of floodplains was dependent on the extent of waterlogging. In addition to organic matter, other reduced substrates (iron sulfide minerals, etc) could be used to promote denitrification in the alluvial aquifer (Appelo and Postma 1993). While the mechanisms are not clear at the present, the pattern in NO3

– concentration at Wollombi Brook relative to water table depth is consistent with observations in other sand aquifers. Other processes Water table evaporation can be a significant loss of water from floodplains in semi-arid climates (Jolly and Walker 1995) and could be a novel mechanism to exchange N between groundwater and the unsaturated zone (Lamontagne et al. 2002b). Water table evaporation occurs when groundwater moves into drying soils by capillary action. It is a complex process that is dependent on climate, the water table depth, and soil properties (Warrick 1988). Water table evaporation could be a mechanism to return some of the dissolved N in groundwater to the unsaturated zone. Thus, the NO3

– pool developing in some shallow floodplain soils could be derived in part from the mineralisation of soil organic matter in situ and in part from the nitrification of NH4

+ originating from groundwater. While water table evaporation probably occurs at Wollombi Brook, its significance to the N cycle is uncertain at the present. Overall, one of the main outcomes of N cycling during a lateral exchange cycle at Wollombi Brook will be to export some of the particulate N deposited out of the stream channel during a flood back to the river as DON and NH4

+. A part of the dissolved N pool in groundwater will be lost during transit in the alluvial aquifer because the numerous redox interfaces encountered will favour the loss of N through denitrification. This tendency to loose N during recycling is consistent with the observed N rather than P limitation of plant growth in semi-arid floodplains (Ogden and Thoms 2002; Ogden et al. 2002). The observations made at Wollombi are also consistent with observations made elsewhere in Australia (Linderfelt and Turner 2001; Smith et al. 2003) that reduced N forms rather than NO3

– can be the principal form of N discharged by groundwater to surface waters. Dissolved nutrients in surface water The Wollombi Brook has been heavily impacted by erosion, with channel incision and the deposition of numerous “sand slugs” occurring throughout the catchment. High input of sediments is usually associated with lower water quality in rivers, including increased

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turbidity and high phosphorus concentrations. However, water quality (in terms of nutrient concentration) in Wollombi Brook was generally excellent. Levels of both dissolved P and dissolved N were generally low under baseflow conditions (Table 2 and Appendix 2). In addition, the principal form of N exported in runoff was DON, which is thought to be a feature characteristic of relatively pristine ecosystems (Hedin et al. 1995; Harris 2001). Overall, nutrient retention at Wollombi Brook is probably facilitated by: 1) extensive beds of charophytes (Kufel and Kufel 2002), 2) the biofilm on the sandy streambed, and 3) hyporheic exchange with the sandy streambed. Charophytes are the only group of freshwater macroalgae and are common in hardwater streams where they form conspicuous mats. Charophytes and the algal biofilms covering the stream bottom actively remove nutrients from the water column and will intercept some of the nutrients discharged through groundwater. By providing an ideal environment for hyporheic exchange, the sandy streambed of Wollombi Brook probably plays a major role in maintaining good water quality at baseflow. Hyporheic exchange with the streambed probably acts as a filter, straining particulate organic matter from the water column (Huettel et al. 1996) and removing nutrients through processes such as denitrification. The potential to completely exchange surface water with the streambed at the scale of hundreds of metres suggests that hyporheic exchange is a significant process in the biogeochemistry of Wollombi Brook.

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Future Research The quantification of nutrient inputs to streams via groundwater remains a challenge. We believe that three major gaps remain: 1) Difficulty to quantify hyporheic exchange and associated biogeochemical processes at the large (i.e., river reach) scale. 2) The confounding effect of hyporheic mixing on nutrient discharge estimates from groundwater. 3) The poor understanding of groundwater–surface water interaction during flood pulse events. These key areas should be considered important avenues for research and are further detailed below. 1) Hyporheic processes measurement and modelling A large body of literature exists on the measurement of hyporheic exchange in streams and rivers (Jones and Mulholland 2000). One promising approach to quantify hyporheic processes at the large scale is through the use of in-stream tracer addition methods (Stream Solute Workshop 1990; Jones and Mulholland 2000), which have not been widely used to date in Australia. On the other hand, in-stream tracer additions are less practical in larger streams and rivers (i.e., often the only ones with some baseflow in semiarid and subtropical Australia) and cannot always differentiate between in-stream and hyporheic retention effects (Fernald et al. 2001). To be practical at larger scales, the use of tracers other than the traditional solutes (i.e., Cl– and Br–) is required. Alternative tracers include dyes such as rhodamine (Fernald et al. 2001) and tritiated water (i.e., water labeled with 3H, a radioisotope of hydrogen; Johansson et al. 2001). These tracers are better at the larger scale because relatively small amounts are required to “label” stream water. However, they also have drawbacks, including retardation with the streambed (dyes) and drinking water quality concerns (dyes and tritiated water). The input of inert dissolved gases (such as Ar, SF6, He, and others) may be a promising alternative for larger scale tracer experiments. Inert gases could be used to label water masses inexpensively and have no water quality impediment. On the other hand, dissolved gases will be gradually lost by exchange with the atmosphere during travel downstream. However, several gases could be used simultaneously to help correct for the loss to the atmosphere (Laursen and Seitzinger 2002). A drawback of in-stream tracer addition experiments is that they cannot always distinguish between in-stream storage and hyporheic effects (Fernald et al. 2001). Thus, the reach-scale information obtained using in-stream tracers should be complemented with detailed measurement of water mixing and nutrient transformations in the hyporheic zone itself. For example, the relatively few porewater profiles collected in this study still provided useful information to understand hyporheic processes in Wollombi Brook. However, because of the large spatial variability (Dent et al. 2001; Fisher et al. 1998), numerous profiles would be require to properly characterise the hyporheic zone in a given stream reach (which could be time consuming and expensive). Thus, there is a need to develop an inexpensive methodology to quantify hyporheic exchange in situ. Temperature may be the ideal “tracer” of hyporheic exchange in porewater profiles (I. Webster, CSIRO Land & Water, personal communication). The daily variation in surface

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water temperature would provide the signal to evaluate hyporheic processes in the subsurface. Temperature can be continuously and inexpensively monitored at a high resolution in both surface water and in the streambed. 2) Integration of groundwater and hyporheic processes Estimates of groundwater nutrient fluxes to streams using traditional approaches (flow nets, etc) should be considered “potential” fluxes because they fail to account for hyporheic mixing in the subsurface (Hinkle et al. 2001). Thus, the assessment of the “actual” groundwater nutrient flux will require the integration of groundwater and hyporheic exchange models. Unfortunately, studies on groundwater and hyporheic processes are seldom integrated (Stanley and Jones 2000). For example, while models exist to predict hyporheic exchange under different flow and geomorphological conditions (Hutchinson and Webster 1998; Packman and Bencala 2000), the impact of groundwater discharge is not considered. Hyporheic and groundwater processes are not independent of one another. For example, field studies have shown that the size of the hyporheic zone is negatively related to the magnitude of groundwater flow (Harvey et al. 1996). A new generation of hyporheic mixing models incorporating groundwater discharge will be required to convert “potential” nutrient fluxes from groundwater to “actual” ones. 3) Groundwater–surface water interactions during flood pulses A pulse of nutrient originating from groundwater could be discharged to rivers during a flood cycle. However, the mechanisms of alluvial groundwater recharge during floods can be quite complex. For example, both vertical diffuse recharge and bank recharge are possible mechanisms of groundwater recharge during floods (Fig. 19). However, which of these mechanisms will be important will be variable from riparian zone to riparian zone. In general, it could be expected that bank recharge would be more significant in V-shaped riverine corridors and diffuse vertical recharge more significant in those with a well-developed floodplain. However, this may not always be the case. For example, at Chowilla, diffuse vertical recharge did not occur in inundated areas of the floodplain because of the presence of sodic clays (Jolly et al. 1994). However, some point vertical recharge was hypothesized to have occurred over limited areas of the floodplain where sandy outcrops occurred (Jolly et al. 1994). Bank recharge also occurred at Chowilla, but the high impedance of the stream banks was found to buffer changes in water table elevation relative to river levels (Narayan et al. 1993; Jolly and Walker 1995). These examples illustrate that the mechanisms of groundwater–surface water interaction during flood pulses could be quite complex and difficult to quantify at the river reach-scale (Jolly et al. 1998). While piezometric monitoring is useful to characterise groundwater–surface water dynamics during flood pulses (Burt et al. 1999; 2002) it is only practical to instrument restricted sections of a riparian zone. Alternatively, hydrograph separation techniques using environmental tracers could be a useful tool to quantify water exchange between streams and alluvial aquifers at the reach scale during flood pulses (Kendall and McDonnell 1998; Cook and Herczeg 1999). With a prior knowledge of the characteristic signature for different environmental tracers for different sources of water, it is possible to determine the contribution of different sources at different stages of a flood event. Once the proportion of groundwater at different stage of the flood is known, estimates of the magnitude of the groundwater nutrient flux could be made based on the measurement of nutrient concentrations in groundwater before and after the flooding event. The blurring effect of hyporheic mixing should be less of a concern early during flow recession, when high groundwater discharge rates would reduce the size of the hyporheic zone even if stream flow is higher (Harvey et al. 1996). Ideally, flooding events should be sampled at

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a high resolution both during the rising and falling limbs of the hydrograph. Information gathered using environmental tracers will be essential to calibrate hydrological models seeking to predict groundwater – surface water interactions during flood cycles.

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Conclusion This study has provided some basic understanding for the significance of groundwater– surface water interactions on nutrient cycles in subtropical alluvial aquifer systems. Some of the key findings of this study include:

• The concentration of some nutrients (NH4+, FRP, Si) and some key elements for

biogeochemical cycles (Fe2+, alkalinity) was elevated in alluvial groundwater and could represent a significant input to surface water.

• The discharge of regional groundwater was not a significant source of nutrients for the brook but could impact on nutrient cycles by changing environmental conditions (i.e. increased salinity, density stratification, anoxia, etc).

• Surface water can be completely exchanged with streambed porewaters at the scale of hundreds of metres in sandbed streams.

• There was a large potential for nutrient transformations during lateral exchange cycles between the stream and the alluvial aquifer because of the presence of several redox interfaces in the alluvial aquifer.

• The depth to the water table in the floodplain may be an important determinant of the redox conditions in the alluvial aquifer.

It is proposed that future research on the significance of groundwater–surface water interactions on nutrient cycles in streams and rivers should focus on:

• The further development of measurement and modelling techniques for hyporheic exchange and associated biogeochemical transformations.

• The integration of groundwater discharge and hyporheic exchange modelling to provide more accurate estimates of nutrient discharge from groundwater.

• The use of hydrograph separation techniques to estimate the contribution of groundwater discharge to stream nutrient budgets during flood pulses.

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Lamontagne, S. Dighton, J. and Ullman, W. 2002a. Estimation of groundwater velocity in riparian zones using point dilution tests. CSIRO Land & Water Technical report 14/02.

Lamontagne, S. Herczeg, A., Leaney, F., Dighton, J., Pritchard, J., Ullman, W., and Jiwan, J. 2001. Nitrogen attenuation by stream riparian zones: Prospects for Australian landscapes. Proceedings of the 8th Murray-Darling Basin groundwater workshop, 4 – 6 September 2001. Victor Harbor, SA.

Lamontagne, S., Leaney, F., and Herczeg, A. 2002b. Streamwater – groundwater interaction: The River Murray at Hattah-Kulkyne Park, Victoria. CSIRO Technical Report 27/02.

Laursen, A.E. and Seitzinger, S.P. 2002. Measurement of denitrification in rivers: an integrated, whole reach approach. Hydrobiologia 485: 67 – 81.

Lawrence, C.R. 1983. Nitrate-rich groundwaters of Australia. Australian Water Resources Council Technical Paper No. 79.

Lawrence, R., Todd, R., Fail, J., Jr., Hendricksen, O. Jr., Leonard, R., Asmussen, L. 1984. Riparian forests as nutrient filters in agricultural watersheds. Bioscience 34: 374 – 377.

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Martí, E., Fischer, S.G., Schade, J.J., and Grimm, N.B. 2000. Flood frequency and stream – riparian linkages in arid lands. pp. 111 – 136. In Jones, J.R. and Mulholland, P.J. [Eds]. Streams and groundwater. Academic Press: New York.

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Appendix 1. Analytical methods

Analyte Method Reference Ca2+, Mg2+, Na+, K+

The samples were determined directly by ICP emission spectrometry. Any concentrations above the linear analytical range were diluted with 1% (v/v) HNO3 before re-analysis.

Standard Methods for the Examination of Water and Wastewater (1999). 20th edition American Public Health Assoc., American Water Works Assoc., Water Environment Federation. Method 3120.

Cl–, SO42– These anions were

determined by ion chromatography (IC) using chemical suppression and electrical conductivity detection.

Standard Methods for the Examination of Water and Wastewater (1999). 20th edition American Public Health Assoc., American Water Works Assoc., Water Environment Federation. Method 4110.

Filterable Reactive Si

The filtered samples were analyzed by segmented flow analysis (SFA) using ammonium molybdate and oxalic acid then reduced with ascorbic acid and determined colorimetrically at 815nm.

Standard Methods for the Examination of Water and Wastewater (1999). 20th edition American Public Health Assoc., American Water Works Assoc., Water Environment Federation. Method 4500-SiO2 Automated method for molybdate-reactive silica (modified). Perstorp Analytical Environmental EnviroFlow 3500 method procedure document 000595 (pers.com.)

Total dissolved Fe (see Ca, Mg, Na & K)

The samples were determined directly by ICP emission spectrometry.

Standard Methods for the Examination of Water and Wastewater (1999). 20th edition American Public Health Assoc., American Water Works Assoc., Water Environment Federation. Method 3120.

NH4+ Determined by

segmented flow analysis (SFA) using the less hazardous sodium salicylate, sodium nitroferricyanide (nitroprusside) and DCIC in alkaline solution with citrate buffer.

Standard Methods for the Examination of Water and Wastewater (1999). 20th edition American Public Health Assoc., American Water Works Assoc., Water Environment Federation. Method 4500-NH3 G Automated phenate method (modified). Krom, M.D. (1980) Spectrophotometric determination of ammonia: a study of a modified Berthelot reaction using salicylate and dichloroisocyanurate. The Analyst, 105, 305-316.

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NO3– + NO2

– Determined by segmented flow analysis (SFA). Nitrate reduced to nitrite by Cu/Cd. Total nitrite then determined colorimetrically after reaction with sulfanilamide and NEDD in acid solution.

Standard Methods for the Examination of Water and Wastewater (1999). 20th edition American Public Health Assoc., American Water Works Assoc., Water Environment Federation. Method 4500-NO3 F Automated cadmium reduction method.

Filterable reactive P

Determined by segmented flow analysis (SFA) using ammonium molybdate and potassium antimony tartrate in the presence of ascorbic acid at pH 1.0 to form a molybdenum blue colour.

Standard Methods for the Examination of Water and Wastewater (1999). 20th edition American Public Health Assoc., American Water Works Assoc., Water Environment Federation. Method 4500-P F Automated ascorbic acid reduction method.

Total dissolved P The samples were determined directly by ICP emission spectrometry. Lower detection limits were achieved using an ultrasonic nebulizer sample introduction system.

Standard Methods for the Examination of Water and Wastewater (1999). 20th edition American Public Health Assoc., American Water Works Assoc., Water Environment Federation. Method 3120.

Dissolved organic carbon

Thermal combustion of filtered solution to form CO2 that is determined by IR detection. Inorganic carbon removed before analysis or determined separately.

Standard Methods for the Examination of Water and Wastewater (1999). 20th edition American Public Health Assoc., American Water Works Assoc., Water Environment Federation. Method 5310 B High-temperature combustion method.

Total dissolved N (TDN) and Dissolved organic N (DON)

TDN was measured by thermal combustion to NO2 and measurement by thermoluminescence. DON was the difference between TDN and (NH4

+ + NO3

-).

SKALAR Analytical B.V. 2000. FormacsHT TOC/TN Analyzer user manual.

Fe2+ Modified phenanthroline method

Lamontagne et al. (2002b)

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Appendix 2. Data summary Appendix 2a. Data summary for the Wollombi piezometer transect, March 2000 sampling trip (in mg L–1 unless otherwise shown).

Site Field Temp (°C)

Field pH

DO

EC (µS/cm)

Lab alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- NH4

+ NO3- FRSi TDFe

River 21.7 6.5 6.9 495 1.35 14 4.4 14 70 110 3.9 - - - <0.1 Wo1A 23.4 6.5 <0.1 485 1.73 20 3.0 7.8 44 62 15 0.30 <0.02 2.5 8.5 Wo1B 24.6 6.4 0.19 461 1.72 16 3.9 11 47 74 1.2 0.42 <0.02 4.6 1.1 Wo2 24.3 6.5 <0.1 470 1.94 23 3.0 9.5 43 62 14 0.24 <0.02 3.9 7.6 Wo3A 22.8 6.4 <0.1 467 1.76 16 2.9 7.9 45 66 9.6 0.22 <0.02 2.6 8.9 Wo3B 24.5 6.4 <0.1 443 1.37 12 3.6 11 46 74 4.5 0.25 <0.02 1.6 9.7 Wo4 21.0 6.4 <0.1 492 0.82 16 3.2 9.3 47 60 63 0.16 <0.02 2.1 11 Wo5 21.7 6.1 ?0.19 417 0.72 15 2.5 11 47 68 48 0.41 <0.02 4.2 2.2

Appendix 2b. Data summary for the Fordwich piezometer transect, March 2000 sampling trip.

Site Field Temp (°C)

Field pH

DO

EC (µS/cm)

Lab alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- NH4

+ NO3- FRSi TDFe

F1A 24.9 6.4 0.20 482 1.92 14 5.9 16 66 90 9.6 0.17 <0.02 3.1 0.30 F1B 22.1 6.4 0.14 578 1.95 16 5.6 17 78 120 1.5 0.12 <0.02 2.9 2.9 F2 25.3 7.0 0.28 509 1.86 14 7.0 15 68 98 6.6 0.20 <0.02 3.5 0.11 F3A 24.6 6.7 0.16 475 1.68 13 6.6 14 66 99 4.5 0.21 <0.02 3.6 <0.1 F3B 22.2 6.6 0.10 654 2.27 16 5.9 19 94 140 1.5 0.19 <0.02 2.5 2.0 F4 25.6 6.5 0.20 545 1.88 14 6.4 16 67 110 6.9 0.25 <0.02 3.0 1.6 F5 24.5 6.5 0.20 314 0.97 11 5.8 11 32 62 11 0.16 <0.02 4.1 <0.1 F6 21.6 6.5 ?0.2 1216 4.21 36 3.0 31 180 240 9.6 0.09 <0.02 8.6 1.8

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Appendix 2c. Data summary for the Warkworth piezometer transect, March 2000 sampling trip.

Site Field Temp (·C)

Field pH

DO

EC (µS/cm)

Lab alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- NH4

+ NO3- FRSi TDFe

River 22.7 7.2 7.90 1867 2.95 29 9.1 44 320 440 123 - - - <0.1 W1A 23.9 6.8 0.1 255 0.81 7.6 2.0 6.1 29 35 17 0.18 <0.02 4.0 3.7 W1B 23.2 6.5 <0.1 316 0.86 11 3.5 11 27 59 10 0.43 <0.02 4.4 1.4 W2A 23.8 6.3 ? 259 0.66 6.7 3.1 4.4 32 40 14 0.19 <0.02 5.0 0.55 W2B 21.9 6.6 0.23 262 0.94 9.3 3.0 8.9 25 45 12 0.30 <0.02 4.3 2.1 W3A 22.9 6.7 ? 263 0.81 8.1 5.0 5.1 28 40 8.7 0.26 <0.02 5.2 0.78 W3B 21.1 6.5 0.40 278 1.26 9.6 3.1 8.5 29 40 1.2 0.37 <0.02 3.6 6.1 W4 21.4 6.1 0.15 235 0.35 7.8 3.7 4.2 23 43 15 0.15 2.1 5.0 7.8 W5 21.2 6.2 ? 524 0.35 16 5.3 14 68 44 36 0.10 2.1 5.0 0.14 Appendix 2d. Data summary for regional bores, Wollombi Brook Valley, March 2000 sampling trip.

Site Field Temp (·C)

Field pH

DO

EC (µS/cm)

Lab alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- FRSi TDFe

Parkes Bore 19.2 6.1 1.2 2230 1.97 41 10 34 450 680 39 6.1 1.9 Hungerford Bore 19.1 6.5 0.1 2370 2.66 47 9.8 69 380 640 90 13 <0.1 Fordwich Bore 18.9 7.1 3.09 1240 6.03 81 0.6 51 120 210 23 24 <0.1 Heathers Bore ----- ----- ----- ----- 2.18 22 8.1 22 160 210 48 16 <0.1 Mulla Villa "spring" 22.6 6.3 0.15? 386 0.97 8.6 4.6 11 49 96 4 1.3 <0.1

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Appendix 2e. Data summary for the Wollombi piezometer transect, May 2000 sampling trip.

Site Field Temp (°C)

Field pH

DO

EC (µS/cm)

Lab alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- NH4

+ NO3- TDP FRSi TDFe

River 8.5 7.7 ? 388 0.97 11 3.9 12 56 96 12 <0.02 0.021 0.01 2.1 0.3 WO1B 9.4 7.7 ? 510 - - - - - - - - - - - - WO1C 10.3 6.6 <0.4 448 1.28 9.4 3 9.3 26 42 12 0.47 0.017 0.02 5.3 0.6 WO2 13 7.3 ? 383 - - - - - - - - - - - - WO3A 14.2 6.9 <0.7 388 - - - - - - - - - - - - WO3B 15.6 6.8 0.4 421 - - - - - - - - - - - - WO4 14.3 6.8 0.8 369 1.22 10 2.5 8.2 41 70 12 0.08 0.019 0.04 4.4 27 WO5 14 6.6 2.4 384 - - - - - - - - - - - - Bore 79056 18 6.1 <0.5 736 1.55 17 4.3 15 93 188 13 1.3 <0.005 0.03 6.8 20 Bore 79055 18.1 6.7 <0.5 2170 15.2 54 22 60 376 328 42 0.45 <0.005 0.04 6.4 <0.1

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Appendix 2f. Water quality summary for the Fordwich transect, May 2000 sampling trip.

Site Field Temp (°C)

Field pH

DO

EC (µS/cm)

Lab alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- NH4

+ NO3- TDP FRSi TDFe

River 9.2 7.5 10.5 510 1.21 14 4.9 16 83 124 7.3 0.03 0.03 0.02 2.2 0.4 F1A 12.4 7.7 <2.5 331 1.62 9.3 3.3 9.6 32 40 7.5 0.28 0.03 0.01 5 0.6 F1B 14.4 7.5 <2.9 539 2.19 15 4.5 19 55 99 2.4 0.23 0.03 0.01 3.8 <0.1 F3A 11.7 7.7 ? 439 1.57 14 5.7 16 57 110 5.3 0.14 0.05 0.01 3.3 0.1 F3B 13.5 7.5 ? 696 2.59 16 5.3 23 88 154 0.7 0.23 0.08 0.01 3.5 0.1 F6 17.1 7.1 ? 2920 0.18 72 4.7 67 413 604 22 0.36 0.01 0.04 9.5 6.3 Bore79057 9.3 7.2 ? 1229 12.7 18 3.2 21 270 74 20 0.80 0.006 0.03 8.1 <0.1 Bore79058 19 7.0 ? 968 0.68 74 1.4 39 95 93 20 0.03 0.44 0.05 23 <0.1

Appendix 2g. Water quality summary for the Warkworth transect, May 2000 sampling trip.

Site Field Temp (°C)

Field pH

DO

EC (µS/cm)

Lab alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- NH4

+ NO3- TDP SrSi TDFe

River 10.5 8.0 ? 643 1.66 13 4.5 17 95 146 17 <0.02 0.009 0.01 2.7 0.3 W1A - - - 457 - - - - - - - - - - - - W1B 15.2 7.9 4.5 312 0.99 9.1 3.2 12 35 71 7.9 0.10 0.016 0.01 3.3 0.2 W2A 14.7 7.8 1.7 236 1.17 8.8 3.8 6.3 23 36 7.5 0.18 0.017 0.01 5.5 14 W2B 16.8 7.9 3.4 286 1.81 19 2.3 8.5 35 64 1.3 0.35 0.039 0.01 4.3 19 W3A 14.7 7.8 1.7 236 0.99 9.6 5.6 7.1 20 41 14.2 0.28 <0.005 0.02 5.6 18 W3B 18.9 7.5 3.2 337 1.37 12 3.6 11 32 64 0.5 0.50 0.007 0.01 5.1 4.8 W4 18.6 7.9 3 187 0.49 5.8 3.5 2.7 22 31 12.6 0.32 0.009 0.01 5 5 W5 19 7.7 <2 258 0.53 5.3 3.2 6.3 34 43 28 <0.02 0.13 0.02 5.1 <0.1 Bore79059 20.2 7.4 <2 7570 10.45 31 14 24 1870 1230 5.2 0.84 0.006 0.04 6 <0.1 Bore79060 19.4 7.3 0.43 1027 0.64 23 5.7 17 306 175 11.9 0.02 0.17 0.03 6.6 0.2

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Appendix 2h. Water quality at several sites in Wollombi Brook, May 2000.

Site Field Temp (°C)

Field pH

DO

EC (µS/cm)

Lab alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- NH4

+ NO3- TDP FRSi TDFe

North Arm. - Milfield Bridge 12.2 6.9 5.2 448 - - - - - - - - - - - - Milfield Bridge 15.5 6.8 - 246 - - - - - - - - - - - - Cedar Creek 12.2 7.0 8.7 462 0.33 11 4.2 12 64 106 17 <0.02 7.0 0.03 3 0.9

Wollombi Site C 13.2 7.3 7.8 426 0.98 11 4 12 57 96 14 <0.02 0.036 0.02 1.9 0.5 Williams Bridge 13.2 - 8.6 373 - - - - - - - - - - - - Paynes Crossing 13.7 6.7 7.8 367 0.86 8.9 4 10 48 86 7.3 <0.02 0.071 <0.01 2.1 0.3

Broke Bridge 15 7.6 9.2 404 0.93 9.2 3.9 11 52 91 6.4 <0.02 0.024 <0.01 1.9 0.3 Cockfighters

Creek 14 7.3 8.5 370 0.88 8.9 4 10 48 84 6.5 0.02 0.016 <0.01 2 0.4 Bulga 12.5 7.7 9.0 536 - - - - - - - - - - - -

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Appendix 2i. Water quality data for Wollombi site, October 2000 trip.

Site Field Temp (°C)

Field pH

DO

EC (µS/cm)

Lab alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- NH4

+ NO3- TDP FRP TDFe

River 19.3 7.0 7.8 - 0.064 11 4.3 12 63 112 9.8 0.04 <0.005 0.012 <0.005 0.08 Wo2 17.9 6.7 0 407 - - - - - 66 - - - - - - Wo3A 17.4 6.7 0.2 406 0.22 - - - - 66 - - - - - - Wo3B 16.7 6.3 1.0 394 0.20 - - - - 68 - - - - - - Wo4 16.5 6.4 <0.1 401 0.17 - - - - 76 - - - - - - Wo5 17.4 6.0 3.2 398 0.04 80 - - - - - - 79055 - - - - - 74 24 64 377 316 42.2 0.79 <0.005 0.109 0.092 1.9 79056 - - - - - 23 4.6 17 105 202 13.6 2.1 <0.005 0.096 0.084 34 WoDP1 - - - - - 26 5.9 22 164 491 3.2 2.2 0.045 0.048 0.008 126 WoDP2 - - - - - 11 2.5 4.3 44 90 2.3 0.97 <0.005 0.021 0.009 48

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Appendix 2j. Water quality data for Warkworth site, October 2000 trip.

Site Field Temp (°C)

Field pH

DO

EC (µS/cm)

Field alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- NH4

+ NO3- TDP FRP TDFe

River 12.7 6.4 9.2 1377 0.272 23 6.8 35 244 343 87.6 0.02 <0.005 0.016 <0.005 0.04 W1B 20.5 6.8 - 254 0.088 - - - - 53 - - - - - - W2A 20.5 6.2 - 217 0.098 - - - - 39 - - - - - - W2B 21.6 6.0 - 251 0.072 - - - - 47 - - - - - - W3A 21.3 6.1 - 228 0.108 - - - - 39 - - - - - - W3B 17.9 6.7 <0.1 274 0.156 - - - - 45 - - - - - - W4 18.2 7.1 <0.1 239 0.074 - - - - 41 - - - - - - W5 20.5 6.7 - 486 0.064 - - - - 108 - - - - - - 79059 - - - - - 34 14 16 2062 1285 0.7 1.4 <0.005 0.17 0.16 0.018 79060 - - - - - 25 5.8 16 237 164 22.8 0.19 1.677 0.038 0.025 0.026 WDP1 23.9 - - 3530 - 67 13 73 649 997 136 1.6 <0.005 0.039 0.011 22 WDP2 24.1 - - 295 - 19 4 12 71 115 30.4 0.23 <0.005 0.019 0.006 2.8 WDP3 - - - 552? - 14 4.5 12 65 80 24.1 <0.02 <0.005 0.018 <0.005 0.3

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Appendix 2k. Water quality data for Fordwich site, October 2000 trip.

Site Field Temp (°C)

Field pH

DO

EC (µS/cm)

Field alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- NH4

+ NO3- TDP FRP TDFe

F1A 15.8 7.1 <0.1 495 0.25 - - - - 113 - - - - - - F1B 17.1 6.8 0.8 545 0.22 - - - - 106 - - - - - - F3A 17.3 6.9 <0.1 497 0.22 - - - - 98 - - - - - - F3B 15.9 7.0 <0.1 629 0.23 - - - - 120 - - - - - - F4 dry dry dry dry dry dry dry dry dry dry dry dry dry dry dry dry F5 16.2 7.1 0.8 229 0.12 - - - - 34 - - - - - - F6 16.9 7.0 - 3230 1.44 - - - - 797 - - - - - -

79057 - - - - - 20 3.1 19 257 76 20.8 0.84 <0.005 0.031 0.017 0.068 79058 - - - - - 68 1 33 90 71 19.8 <0.02 0.029 0.037 0.022 0.025 FoDP1 - - - 1425 - 54 2.4 27 219 292 3.1 0.47 <0.005 0.088 0.06 5.5 FoDP2 - - - 1215 - 26 5.5 21 207 271 1.5 1.1 <0.005 0.029 0.009 19 FoDP3 - - - 958 - 37 3 21 127 210 1.8 0.56 <0.005 0.04 0.014 1.9

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 50

Appendix 2l. Water quality data for the Wollombi site, March 2001.

Site Field Temp (·C)

DO

EC (µS/ cm)

Lab alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- NH4

+ NO3- TDP FRP TDFe Fe2+ DOC FR

Si

River - - 394 1.00 10 5.3 10 49 86 9.7 0.28 0.19 0.012 <0.005 0.9 - - 3.8 Wo2 21.6 0.11 558 1.51 19 2.1 9.1 53 100 <0.4 1.1 0.049 0.038 0.005 30 n.s. - 5.1 Wo3B 22 0.1 464 1.08 12 2.3 10 41 87 0.7 0.84 0.059 0.19 0.127 26 30 - 4.6 Wo4 21 0.1 530 1.33 15 1.7 15 41 93 5.4 0.50 0.098 0.063 0.029 34 39 - 5.3 Wo5 20.2 0.25 418 0.56 5.9 1.6 8.9 47 85 16 0.98 0.067 0.092 0.052 17 20 - 4.4 79055 19.6 0.15 2370 16.1 75 26 67 390 330 25 1.8 0.067 0.21 0.132 2.7? 14 2.7 7.9 79056 19.2 0.1 934 1.63 20 3.5 20 100 200 6.2 2.2 0.10 0.22 0.183 44 20 2.6 6.5 WoDP1 - - - 0.93 10 3.9 11 50 200 8.6 0.25 0.16 0.039 0.011 1.5 1.7 - 3.6 WoDP2 20.9 - 265 0.92 15 2.9 5.1 37 59 3.5 1.8 0.063 0.043 <0.005 27 27 - 4.3

Appendix 2m. Water quality for the Warkworth site, March 2001

Site Field Temp (·C)

DO

EC (µS/ cm)

Lab alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- NH4

+ NO3- TDP FRP TDFe Fe2+ DOC FR

Si

River 22.2 - 484 - 11 5.8 13 67 110 13 0.19 0.1 <0.005 <0.005 0.5 - 8.4 3.9 W1B 23.8 0.9? 616 1.19 - - - - 100 29 - - - - - 1.1 - - W2B 25.3 0.1 474 1.38 - - - - 51 26 - - - - - 5.8 3.3 - W3A 23.5 0.1 304 1.20 4.9 3.8 3.7 43 42 14 0.24 0.07 0.018 <0.005 11 10.9 6.4 5.6 W3B 22.8 0.1 309 1.40 10 3.5 9.4 32 49 2.3 1.0 0.04 0.029 0.009 8.5 8.5 2.8 5.1 W4 25.1 0.1 155 0.62 4.7 3.6 2.5 17 13 17 0.54 0.03 <0.005 <0.005 6.3 6.8 2.5 5.0 W5 22.4 1.8 175 0.55 2.3 2.5 2.6 27 34 9.5 0.21 0.21 0.008 <0.005 0.1 0.12 4.2 4.7

79059 22.9 0.22 6620 11.7 20 15 11 2100 920 <0.4 0.45 0.04 0.25 0.14 <0.1 0.19 1.5 6.6 79060 22 0.21 1712 8.72 17 6.7 11 390 180 12 0.37 0.16 0.043 0.012 <0.1 <0.1 2.7 6.2 WDP1 22.3 - 827 2.95 21 5.7 22 150 90 9.6 0.98 0.008 0.025 <0.005 5.3 5.9 - 4.8 WDP2 22.9 - 334 1.04 13 4.5 10 31 49 22 0.44 0.01 0.047 0.023 5.3 - - 5.6 WDP3 21.5 - 522 2.38 11 5.8 13 67 110 13 0.19 0.1 <0.005 <0.005 0.5 - - -

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Appendix 2n. Water quality data at the Fordwich site, March 2001

Site DO

EC (µS/ cm)

Lab alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- NH4

+ NO3- TDP FRP TDFe Fe2+ DOC FR

Si

River 7.4 445 1.07 10 5.4 12 56 100 7.5 <0.02 0.084 0.015 <0.005 0.5 0.3 6.8 3.6 F1B 0 480 1.69 12 5.4 15 60 100 2.8 0.19 0.008 0.048 0.018 2.8 2.9 3.3 3.8 F3A 0.1 341 1.21 10 4.5 11 40 70 8.2 0.25 0.021 0.017 <0.005 0.9 0.93 5.2 3.7 F3B 0.07 709 3.02 18 5.4 25 81 130 <0.4 0.33 0.002 0.080 0.045 12 11 4.8 4.0 F5 0.2 229 1.77 6.4 3.6 5.5 32 16 5.9 0.17 <0.005 0.022 <0.005 2.5 3 2.9 4.1 F6 1.4 1955 7.26 49 1.8 48 300 380 47 0.22 0.009 0.093 0.032 26 22 11.5 12

79057 0.1 1236 - 69 1.3 35 92 74 18 <0.02 0.076 0.049 0.011 <0.1 - - 22 79058 0.3 954 8.47 - - - - - - - - - - - - 1.3 - FoDP1 - 1478 6.60 54 3.2 29 230 320 <0.4 0.54 <0.005 0.031 <0.005 1.9 3.1 - 15 FoDP2 - 1309 4.34 25 6.4 21 210 280 <0.4 0.98 0.019 0.024 <0.005 16 19.4 - 6.5 FoDP3 - 585 1.79 16 6.5 16 73 130 5.9 0.38 0.083 0.014 <0.005 0.6 0.6 - 4.4

Appendix 2o. Water quality data at the Wollombi site, November 2001.

Site Field Temp (°C)

pH

EC (µS/ cm)

Field alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- NH4

+ NO3- TDP FRP TDFe Fe2+ DOC FR

Si

River - - - 0.98 - - - - 125 - - - - - - 0.2 - - Wo2 20.4 6.9 470 1.88 12 2.2 8.3 41 76 - 0.28 0.067 0.023 0.023 31 - 4.1 7.7

Wo3B 18.7 6.6 463 1.8 9.4 2.5 9.6 43 84 - 0.48 0.02 0.255 0.221 23 - 3.5 6.7 Wo4 19.5 6.6 458 1.76 8.9 2.5 10 39 84 - 0.17 0.019 0.048 0.049 27 - 3.0 7.3 Wo5 18 6.2 395 1.12 4.8 1.9 8.4 45 84 - 0.40 0.025 0.293 0.258 15 - 3.7 7.1

79055 18.8 6.9 2410 14.8 64 23 61 347 333 - 1.5 0.019 0.242 0.197 1.8 2.1 6.7 8.2 79056 19.1 6.5 1088 7.98 20 4.7 22 108 228 - 2.1 0.015 0.305 0.264 54 40 2.9 8.3

WoDP2 17.7 6.6 465 2.1 6.2 2.4 4 41 81 0.5 1.2 0.02 0.475 0.474 54 44 - 6.5

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Appendix 2p. Water quality data for the Warkworth site, November 2001. Site Field

Temp (·C)

pH

EC (µS/ cm)

Field alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- NH4

+ NO3- TDP FRP TDFe Fe2+ DOC FR

Si

River - - - 2.5 19.5 6.2 30 197 315 67 0.055 <0.005 0.012 <0.005 <0.1 <0.1 4.5 2.9 W1B - - - 0.74 5 2.7 6.4 31 62.2 10 0.20 <0.005 0.008 0.005 0.1 0.12 3.1 6.1 W2B - - - 0.74 7.6 3.3 6.4 34 66.3 6.9 0.35 0.012 0.01 <0.005 0.7 0.92 2.5 5.9 W3A 25 6.7 320 0.46 6.5 3.1 2.9 39 75.4 17 0.28 0.075 0.008 <0.005 16 17 3.2 5.8 W3B 22.1 7.0 291 0.9 6.9 2.8 7.1 30 70.9 1.4 0.54 0.011 0.021 0.015 4.8 5.6 2.5 6 W4 6.6 288 1.0 7.3 3.6 3 35 55.5 18 0.22 0.021 0.016 0.01 12 14 3.1 6.3 W5 24.7 6.4 530 1 12 4.7 14 60 103 35 0.065 4.6 0.027 0.021 <0.1 <0.01 5.3 5.6 79059 22.5 7.6 5910 37 11 9.7 8.2 1583 220 0.9 1.2 <0.005 0.32 0.24 <0.1 0.02 - 6.8 79060 23.8 7.2 1874 11 22 6.4 17 392 216 8.7 0.27 0.47 0.033 0.036 <0.1 <0.01 1.2 7.4 WDP1 - - - 2.7 12 3.5 13 77 113 0.7 0.50 0.019 0.018 0.012 6.6 8.0 - 5.9 WDP2 - 6.8 268 0.7 7.5 4.7 4.9 34 66 11 0.15 0.023 0.027 0.024 3.8 4.4 - 6.9 WDP3 23.8 7.0 468 1.8 15 4.4 13 68 81 19 0.12 0.036 0.013 0.012 0.3 0.5 - 7.4

Appendix 2q. Water quality data for the Fordwich site, November 2001.

Site Field Temp (·C)

pH

EC (µS/ cm)

Field alk.

(meq/L)

Ca2+ K+ Mg2+ Na+ Cl- SO42- NH4

+ NO3- TDP FRP TDFe Fe2+ DOC FR

Si

River - - - 2 28 6.4 38 213 420 16 0.043 0.009 0.01 <0.005 <0.1 0.23 4.9 2.7 F1B 18.1 7.0 633 2.0 15 4.1 18 71 136 <0.4 0.42 0.012 0.033 0.022 6.7 8.2 4.4 4.6 F3A 19.5 7.0 566 1.9 13 4.1 16 67 117 <0.4 0.33 0.01 0.014 0.006 1.5 1.9 4.7 4.3 F3B 18.3 7.0 814 2.5 18 4.8 23 104 170 <0.4 0.38 0.007 0.072 0.069 11 13 4.8 4.8 F5 19.2 6.8 316 1.44 10 3.7 7 33 64 3 0.17 0.015 0.012 0.008 5.1 5.6 5.1 4.9 F6 19.4 6.8 6430 14.7 27 4.8 24 79 1740 122 0.74 <0.005 0.03 0.035 - 42 13.5 14

79057 20.9 6.1 973 9.5 19 2.9 20 249 81 - 0.98 0.01 0.017 0.018 <0.1 0.076 1.0 9 79058 20.3 7.2 879 6.6 69 0.9 35 91 102 - 0.075 0.44 0.022 0.025 <0.1 <0.01 <0.5 25 FoDP1 - - - 6.0 51 2.8 28 213 310 <0.4 0.63 0.07 0.027 0.021 4.1 5 - 14 FoDP2 - 6.8 1276 3.8 21 6 19 188 262 2.8 1.3 0.022 0.091 0.092 17 19 - 7 FoDP3 18.8 7.1 1222 3.4 43 3.9 31 145 270 0.5 0.61 0.028 0.035 0.036 4.7 5.5 - 10

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Appendix 2r. Water quality data for mini-piezometer bundles (BP) and stream sediment profilers (R) at Warkworth, April 2001. Piezo

EC Ca2+ K+ Mg2+ Na+ Cl- SO4

2- NH4+ NO3

- TDP FRP TDFe Fe2+ FRSi

BP1-4 165 2.4 2.6 24 2.7 33 9.1 <0.02 0.090 0.007 <0.005 <0.1 <0.1 4.3 BP1-5 276 5.3 7.6 28 3.5 60 6.3 <0.02 0.007 0.006 <0.005 <0.1 <0.1 5.2 BP1-6 473 15 19 48 4.9 82 26 <0.02 <0.005 0.012 <0.005 <0.1 <0.1 5.4 BP1-7 492 15 15 57 5.2 87 23 0.06 0.036 0.017 <0.005 0.1 <0.1 5.7 BP1-8 484 15 15 60 5.1 89 24 0.02 0.006 0.011 <0.005 <0.1 <0.1 5.8

BP2-3 219 - - - - - - - - - - - <0.1 - BP2-4 163 2.0 2.2 25 2.8 33 10 0.04 0.14 0.006 <0.005 <0.1 <0.1 4.4 BP2-5 156 1.6 1.7 24 2.3 31 8.4 <0.02 0.24 0.008 <0.005 <0.1 <0.1 4.3 BP2-6 clogged - - - - - - - - - - - - - BP2-7 clogged - - - - - - - - - - - - - BP2-8 395 11 9.7 45 4.7 65 18 <0.02 1.5 0.022 <0.005 <0.1 <0.1 5.9

BP3-4 208 7.3 7.7 33 4.8 47 37 0.05 0.11 <0.005 <0.005 0.1 <0.1 5.5 BP3-5 303 11 12 42 5.1 72 35 <0.02 5.6 0.006 <0.005 <0.1 <0.1 5.5 BP3-6 352 12 12 45 5.1 82 33 <0.02 5.2 0.006 <0.005 <0.1 <0.1 5.4 BP3-7 254 6.3 6.2 30 4.0 27 6.1 0.12 2.2 0.006 <0.005 <0.1 <0.1 4.9 BP3-8 172 3.9 3.9 24 3.1 42 11 0.03 0.25 0.032 <0.005 <0.1 <0.1 4.8 BP3-9 175 3.9 4.0 23 3.2 38 15 <0.02 0.22 0.052 <0.005 <0.1 0.16 5.3

R1-10 - 13 13 53 3.4 98 18 0.38 0.01 0.027 <0.005 6.3 6 4.9 R1-80 - 19 24 49 5.0 140 35 0.60 0.037 0.039 <0.005 12 12 5.0

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Appendix 2s. Water quality data in mini-piezometer bundles and stream sediment profiles (R) at Warkworth, November 2001. Piezo

Ca2+ K+ Mg2+ Na+ Cl- SO4

2- NH4+ NO3

- TDP FRP TDFe Fe2+ FRSi

BP1-4 9 4.9 10 50 95 23 0.12 1.1 0.011 0.005 <0.1 <0.01 5.9 BP1-5 12 5.1 16 53 110 31 0.04 3.8 0.009 0.007 <0.1 <0.01 6.5 BP1-6 17 5.8 16 59 110 36 0.13 0.07 0.007 <0.005 <0.1? 3.6 6.6 BP1-7 12 4.1 14 56 96 28 0.06 0.57 0.037 0.007 <0.1 0.17 6.9 BP1-8 11 3.5 13 56 86 26 0.05 0.44 0.015 0.007 <0.1 <0.01 7.1

BP2-7 10 5.6 16 59 100 39 <0.02 4.6 0.01 0.006 <0.1 <0.01 5.4 BP2-8 15 5.5 14 58 85 32 0.03 4.6 0.011 0.006 <0.1 <0.01 7.3

BP3-6 15 5.3 16 42 110 30 0.02 2.1 0.008 <0.005 <0.1 <0.01 6 BP3-7 15 5.4 15 48 110 33 <0.02 3.7 0.008 <0.005 <0.1 <0.01 6 BP3-8 15 6.1 15 53 110 37 <0.02 4.0 0.006 <0.005 <0.1 <0.01 5.3 BP3-9 14 6.1 14 55 110 36 0.02 4.6 0.011 0.01 <0.1 <0.01 5.7

R3-80 5.1 3.7 7.1 124 - - 0.40 <0.005 0.028 0.022 1.7 2.1 4.5 R4-20 17 7 22 139 - - 0.34 0.05 0.015 0.012 0.8 0.94 4.8 R4-75 29 9 51 412 - - 0.70 0.02 0.012 0.008 10 12 5

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Appendix 2t. Water quality data in mini-piezometer bundles at the Wollombi site, October 2000. Depth is in cm below the water table.

Piezo Depth Ca2+ K+ Mg2+ Na+ Cl- SO42- TDFe TDP FRP NH4

+ NO3-

BP1-3 10 31 5.8 26 39 65 3.9 0.8 0.033 0.008 <0.02 <0.005BP1-4 30 29 5.6 24 39 61 2.9 1.0 0.029 0.010 0.10 <0.005BP1-5 50 22 5.6 21 37 57 6.9 2.1 0.051 0.031 0.15 <0.005BP1-6 75 16 4.8 16 36 56 6.2 9.8 0.039 0.025 0.21 <0.005BP1-7 100 16 4.6 14 37 54 5.9 11 0.052 0.034 0.22 <0.005BP2-3 10 8.1 2.4 6.0 35 55 5.8 26 0.038 0.019 0.16 <0.005BP2-4 30 7.3 1.8 7.7 35 56 10.3 18 0.048 0.036 0.08 <0.005BP2-5 50 6.2 1.6 5.9 36 59 8.2 20 0.102 0.086 0.02 <0.005BP2-6 75 5.2 1.7 6.8 37 61 7.7 13 0.178 0.145 0.16 <0.005

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Appendix 3. Some useful tools for groundwater–surface water interaction studies A number of tools were developed to complete the sampling for water quality using the standard piezometers originally installed at Wollombi Brook. These additional tools were needed to collect more detailed groundwater and sediment porewater nutrient profiles and to sample at a greater depth within the alluvial aquifer. These tools included bundles of mini-piezometers, drive points, and a stream porewater profiler. These are described in further detail below. Mini piezometer bundles Mini-piezometer bundles were used to collect detailed groundwater profiles in the alluvial aquifer. These consisted of various lengths, cut to the required depth interval, of 6 mm nylon pressure tubing firmly attached to a length of 12mm PVC electrical conduit, with nylon cable ties. To prevent clogging, the end of the nylon tubing was covered with mesh cloth held in place with stainless steel wire. A PVC spear point was glued to the base of the 12 mm pipe. The pipe can be slotted for use as a water level recording well. Mini-piezometers were sampled with a hand-held peristaltic pump (Cole Palmer) or a small 12V vacuum pump system. While mini-piezometer bundles can be installed using jet-coring techniques, they were installed using excavation within a casing at Wollombi Brook to prevent the introduction of foreign water to the aquifer. Clogging was frequently encountered following installation but could generally be solved by vigorous pumping. Mini-piezometer bundles are inexpensive and allow the collection of water samples with a greater resolution than with more traditional designs. The main drawback of mini-piezometer bundles is that only small amounts of water can be withdrawn and that they have a tendency to clog when silts and clays are present.

Figure A3.1: Design of mini-piezometer bundles.

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Pumping test To accurately measure gradients in element concentration in a groundwater profile, the sphere of influence of a mini-piezometer must be maintained as small as possible. A simple pumping test was made at Fordwich in October 2000 assess how the volume of groundwater extracted could influence concentration measurements with a mini-piezometers. One mini-piezometer located 20 cm below the water table (i.e. where a steep gradient in nutrient concentration was present) was continuously pumped and samples collected every 150 mL for up to 1050 mL. The volume of groundwater extracted greatly exceeded the volume normally collected during regular sampling events (<150 mL). There was a small decrease in TDFe and TDP concentration and a small increase in NO3

– concentration during the course of the pumping test. NH4+ concentrations varied

more widely but were near the detection limit of the analytical method used (Fig. A3.2). These results suggest that as the pumping test proceeded, the volume of aquifer sampled increased and groundwater with different element concentration was sampled. Thus, the element concentration measured will be a function of how much water is pumped. However, this artifact appears negligible under normal sampling procedures (<150 mL). This volume corresponds to a radius of influence of ~5 cm for an aquifer porosity of 0.3.

Figure A3.2: Pumping test for a mini-piezometer located near the water table.

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Contamination of nylon tubing During a leaching experiment, it was noted that the nylon tubing used to build the mini-piezometers produced significant concentrations of organic C, S, and N (Table A3.1). While the amount of dissolved organic matter produced may decrease as the tubing ages, it is not recommended to use nylon tubing for the mini-piezometer bundles if water samples are to be analysed for dissolved organic matter.

Table A3.1: Dissolved organic matter leaching experiments from nylon tubing. All experiments made with one-metre lengths of new nylon tubing rinsed with three volumes of distilled water. One experiment was made by gently pumping distilled water through the tubing and by collecting samples at regular intervals. In a second experiment, distilled water was left to incubate for 24 hours in both nylon tubing and the sampling bottles used to collect samples in the field (mean and standard deviation for three replicates for each).

DOC (mg L–1)

DON (mg L–1)

DOS (mg L–1)

a) Continuous flow 50 mL 7.5 1.1 2.6 100 mL 6.4 0.8 1.3 200 mL 3.9 0.5 0.8 b) 24 h incubation Nylon tubing 447 ± 3 49.7 ± 0.6 77.0 ± 4 Poly. bottles 0.57 ± 0.06 <0.05 <0.1

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Drive Point The installation of piezometers by excavation within a casing was not practical below a few metres in the Wollombi Brook alluvial aquifer because a sand slurry tended to rapidly fill the inside of the casing from the bottom. Riparian landowners frequently insert “spear points” in the alluvial aquifer using jet coring techniques, but this has the drawback of introducing large amounts of water to the aquifer. Deep (>4 metres) piezometers were inserted in the alluvial aquifer by using a slightly modified version of the drive point systems designed by P. Peter and colleagues (CSIRO Land & Water, Adelaide). The drive points consist of one-meter lengths of 12 mm ID PVC or stainless steel tubing (Fig. A3.3). The distal end consists of a 5 cm fritted stainless steel filter and driving tip. Mesh size for the filters was 200 µm. Additional lengths of tubing can be joined using PVC connectors and glue. The drive points are installed by percussion using 12 mm stainless steel rods. The rods are designed to fit snuggly within the drive points to prevent breakage during installation. A blunt ended rod was used during installation. However, it is also possible to retrieve the drive point using a fitted rod tip. Drive points are installed by gently hammering with a sledgehammer on a brass shoe. Drive points were inserted in such fashion down to six metres in wet sands at Wollombi and elsewhere. The drive points were sampled by inserting a length of 6 mm nylon tubing and a hand pump. A sampling device using displacement with an inert gas is also available (A. Nefiodovas, CSIRO L&W, personal communication)

Figure A3.3: Drive point piezometer.

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Porewater profiler Several samplers have been recently proposed for measuring the porewater chemistry of sandy streambed sediments at the centimetre scale (Duff et al. 1998; Berg and McGlathery 2001). These were tried at Wollombi Brook but were felt inadequate for characterising the mixing of groundwater and surface water in the hyporheic zone in different parts of the stream (which occurs at the scale of tens of centimetres). Smaller-scale porewater samplers were also susceptible to clogging, which considerably increased the time required to collect the porewater profiles. A stream porewater profiler was designed to complement the information gathered with finer scale profilers (Fig A3.4). The stream profiler consisted of a 1.2 m length of 12 mm 316 stainless steel tubing. A stainless steel spear point and screen assembly was welded at the distal end and covered with a 500 µm Nitex mesh. A length of 6 mm nylon tube was inserted into the stainless steel tube down to the screened section and sealed in place with a rubber-potting compound. To aid in the insertion of the tube into the sediments, a brass fitting was constructed by welding the body of a 2.5 cm brass stopcock to the top of the tube. The profiler was pushed in the streambed by hand or by gentle hammering. Samples were usually collected every 10 cm by gently pumping with a hand pump. While the stream profiler cannot be used to measure potential changes in nutrient concentration occurring at the cm-scale (such as the ones at the benthic nitrification–denitrification interface), it was simple and robust to use in the field and the appropriate size to characterize hyporheic mixing at Wollombi Brook. The profiler was also useful to measured hydraulic head at depth relative to surface water using a potentiomanometer.

Figure A3.4: Streambed porewater profiler.

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 61

Appendix 4. Parsons Creek tracer addition The addition of conservative tracers in-stream is a powerful method to estimate the flux of groundwater along stream reaches (Triska et al. 1993). The method involves the addition of a tracer at a constant rate and the measurement of the changes in tracer concentration at different points along the stream section. The test must run long enough for the salt concentration to reach a plateau at all stations. If groundwater influx occurs between measuring stations, the plateau tracer concentration will be diluted in downstream stations relative to the upstream ones. This dilution of the tracer can be used to estimate the groundwater discharge between stations. Furthermore, the shape of the tracer concentration versus time curve can be use to quantify hyporheic exchange parametres (Harvey and Wagner 2000). To complement the existing piezometric work, an in-stream tracer addition was considered for Wollombi Brook. However, to test the applicability of such tracer addition on the relatively large Wollombi Brook, a preliminary test was performed on a smaller tributary (Parsons Creek, near the town of Broke). Aside from a smaller size, Parsons Creek is similar to Wollombi Brook (i.e., it is a sand-bed stream). The groundwater discharge estimated with the in-stream tracer addition was also compared to the one inferred from Darcy fluxes using potentiomanometer measurements along the streambed. Methods In-stream tracer addition A 200 m section of Parsons Creek was selected for the experiment. At the study site, Parsons Creek is a gently meandering riffle and pool stream. At the time of the experiment (3 April 2001), stream width was approximately 5 metres and flow rates <0.05 m3 s-1. A dripper was installed and measuring stations set-up downstream at 31, 82, and 138 metres (Stations 1, 2, and 3). The dripper consisted a garden irrigation system installed perpendicularly across and 50 cm above the stream. The tracers (105 g L-1 Br–) was added using a peristaltic pump (Cole-Parmer). Sprinklers were placed at regular intervals on the irrigation tubing to ensure that tracer was sprayed across the whole channel width. The pumping rate remained constant during the test at 2.0 L min-1, as verified by regular flow rate measurements (time to empty a 1 L graduated cylinder). Samples were collected using an ISCO sampler at stations 1 and 2 and manual samples at Station 3. Bromide concentration was measured at the Analytical Chemistry Laboratory at CSIRO Land & Water, Adelaide. At each measuring station, stream discharge (Qo) was estimated from the in-stream tracer addition using:

QCCCCQ l

bP

blo

⋅−

−= (A4.1),

where Cl is the concentration of the added tracer, Cb the background tracer concentration, Cp the tracer concentration at plateau phase, and Ql the dripping rate of

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Groundwater – surface water interactions at Wollombi Brook: Biogeochemical processes 62

the tracer. The difference in discharge between downstream and upstream stations is an estimate of the groundwater inflow rate. Potentiomanometer measurements Immediately following the test, one hydraulic gradient measurement was made at mid-stream every 12.5 metres from Station 1 to Station 3. The description and use of potentiomanometres has been made elsewhere (Winter et al. 1988). Briefly, a stainless steel drive point was inserted one meter into the streambed. The difference in hydraulic head between the drive point and the stream was measured with the potentiomanometer. These hydraulic gradients were combined with estimates of the hydraulic conductivity of the streambed and of the streambed surface area to evaluate the groundwater flux between stations (see below). To estimate the streambed surface area, cross-sectional profiles were measured every 25 metres from Station 1 to 3. Groundwater discharge between stations was estimated using:

AiKQG= (A4.2),

where QG is the groundwater discharge rate (m3 min-1), K the hydraulic conductivity of the streambed (m min-1), i the hydraulic gradient, and A the streambed surface area (m2). Results Following one hour of tracer addition, the plateaus in Br– concentration were 111.3, 108.8 and 105.0 mg L-1 at stations 1, 2, and 3, respectively (Fig. A4.1). This dilution of the tracer in downstream stations indicates that groundwater inflow occurred between stations. The estimated stream discharge rates were 1928, 1972, and 2044 L min-1 for stations 1, 2, and 3, respectively. Therefore, the flux of groundwater was ~44 L min-1 between stations 1 and 2 and ~72 L min-1 between stations 2 and 3.

Figure A4.1: In-stream tracer addition test at Parsons Creek, 3 April 2001. CP1, CP2, and CP3 represent the plateaus in tracer concentration at stations 1, 2, and 3, respectively. Tracer addition started at 10:40 and was terminated at 11:33.

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Darcy velocities Significant variability in hydraulic head was found between stations 1 to 3 (Fig. A4.2). In general, hydraulic gradients were positive (towards the surface), with the exception of flat gradients in a pool between stations 2 and 3. Additional measurements were made along the stream edge at station 1, where conspicuous accumulations of Fe oxides were visible. Hydraulic gradients were two to three times larger near the Fe flocculates, suggesting a greater groundwater discharge near the stream edge than at mid-stream. Figure A4.2: Variations in hydraulic head at mid-stream between stations 1 and 3. The hydraulic conductivity of the streambed sands was not known but was estimated as ranging between 0.01 to 0.1 m min-1 (clean sand; Freeze and Cherry 1979). Using this range in K values, the groundwater inflow rate would be between 6.7 to 67 L min-1 between stations 1 and 2 and 5.9 to 59 L min-1 between stations 2 and 3. Because the possibility of higher hydraulic gradients near the stream edge was neglected, these Darcy-derived groundwater discharge rates are probably conservative. Overall, the groundwater discharge rates obtained with the in-stream tracer and the Darcy velocity approaches are qualitatively similar (i.e., of the same order of magnitude).

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Table A4.1: Groundwater discharge along two segments of Parsons Creek based on potentiomanometer measurements.

Section Hydraulic gradient

Surface area (m2)

Estimated K (m min-1)

QG (L min-1)

Stations 1 to 2 0.003 223 0.01 6.7 0.003 223 0.03 20 0.003 223 0.1 67 Stations 2 to 3 0.0027 219 0.01 5.9 0.0027 219 0.03 18 0.0027 219 0.1 59 Discussion In-stream tracer addition is a potentially very accurate method to estimate groundwater discharge along stream segments but it has limitations. The experimental set-up must be carefully designed to accommodate the stream discharge rate and the expected groundwater flow rate. In our case, it would have been preferable to at least double the distance between measuring stations and double or triple the duration of the tracer addition to ensure a clearer difference in plateau concentration between stations. The mixing distance between the dripper and the first station should also have been greater because there was some evidence that the tracer was not perfectly mixed across the stream channel at Station 1. Based on the results from Parsons Creek, a similar in-stream tracer addition would be possible but logistically difficult at Wollombi Brook (where flow rates can be above 1 m3 s–1). Distance between stations would have to be in the hundreds of metres and the tracer would have to be added for several days (because of delayed transit in larger pools). Approximately 20 kg of Br– were added to Parsons Creek during the field trial. Thus, several hundred kilograms of Br– would be required at Wollombi Brook for a similar experiment.


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