nutrient limitation and interactions with organic matter and … · nutrient limitation and...

Nutrient limitation and interactions with

organic matter and sediments within dryland

streams of the Pilbara region of northwest

Australia

Jordan Andrew Iles

B.Sc. (Environmental Biology) University of Technology Sydney

This thesis is presented for the degree of Doctor of Philosophy

of The University of Western Australia

School of Biological Sciences

2019

ii

THESIS DECLARATION

I, Jordan Andrew Iles, certify that:

This thesis has been substantially accomplished during enrolment in the

degree.

This thesis does not contain material which has been submitted for the award

of any other degree or diploma in my name, in any university or other tertiary

institution.

No part of this work will, in the future, be used in a submission in my name,

for any other degree or diploma in any university or other tertiary institution

without the prior approval of The University of Western Australia and where

applicable, any partner institution responsible for the joint-award of this

degree.

This thesis does not contain any material previously published or written by

another person, except where due reference has been made in the text and,

where relevant, in the Declaration that follows.

The work(s) are not in any way a violation or infringement of any copyright,

trademark, patent, or other rights whatsoever of any person.

This thesis does not contain work that I have published, nor work under

review for publication.

Signature:

Date: 31th

August 2018

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ABSTRACT

This thesis seeks to increase understanding of the ecological functioning of

ephemeral streams of the arid Pilbara region of northwest Australia. In this sub-

tropical region, nutrient and organic matter dynamics are closely coupled to the

highly episodic flows generated by cyclonic recharge, which are in turn punctuated

by prolonged periods of drought. Flood events redistribute nutrients and organic

matter throughout the catchment and into streams, but these events tend to be

irregular in occurrence as well as short-lived. Hence, the typical state of surface

water of these streams for the majority of the year – and sometimes for multiple

consecutive years – are as contracted and isolated pools. The nutrient status and

metabolism of Pilbara streams are thus likely to be strongly influenced by both

evaporative loss and groundwater connectivity of individual pools. In this thesis, I

investigated: i) patterns in biogeochemical processes in pools of varying hydrologic

connectivity across one of the largest river catchments in the Pilbara; ii) phosphorus

adsorption and desorption characteristics of stream sediments and how organic

matter interacts with sorption processes; and iii) ecosystem metabolism and the

response of phytoplankton, periphyton, and charophytes to increased nutrient

availability in persistent versus more ephemeral pools.

My analysis of the biogeochemical nature of surface waters suggests that the

intermittent and ephemeral streams within the Fortescue River catchment can be split

into three broad groups; spring-fed streams, streams with connectivity to alluvial

water, and ephemeral streams disconnected from alluvial water. I observed that

nutrient concentrations, water isotope composition, as well as dissolved organic

iv

matter (DOM) concentration and composition were strongly differentiated between

the upper and lower sub-catchments of the Fortescue River. I thus sought to explore

further some of the longitudinal in-stream processes that affect the biogeochemistry

of ephemeral and intermittent streams, particularly in the confined upper gorges,

where many streams are dominated by iron-rich sediments but may also contain

considerable amounts of calcrete.

Preliminary observations indicated that these streams act as a sink for allochthonous

organic matter due to the accumulations of leaf litter within the pools. Leaching of

this material in situ contributes to the total loading of dissolved organic matter

(DOM) in the system, with a presumed concentration of DOM compounds occurring

at the litter-sediment-water interface. In order to understand the mechanisms that

might explain the low concentrations of phosphorus (P) in the water column, I

investigated interactions between inorganic phosphate, dissolved organic matter

(DOM) and iron-rich stream sediments (hematite and goethite 39-50%) by

characterising adsorption-desorption kinetics. Sediment adsorption of P closely

follows Freundlich and Langmuir isotherm models with Langmuir P sorption

maxima ranging from 0.106 to 0.152 mg g-1

. P sorption characteristics did not differ

among pools of contrasting hydrological connectivity but were altered by DOM

additions. While moderate DOM additions (~5 mg L-1

DOC) from leaf litter

leachates reduced sediment P adsorption capacity, more concentrated additions (~50

mg L-1

DOC) likely saturated sediment surface adsorption sites and produced P-OM-

Fe complexes, resulting in removal of phosphate from solution. There was a

preferential sorption of high spectral slope OM to sediments when phosphate was at

negligible to low concentrations. Increasing the concentration of DOM in solution

v

also increased the amount of P initially desorbed from sediment. These findings

reveal that interactions of phosphates with organic matter inputs, often from

allochthonous sources, may be important in regulating nutrient availability in both

ephemeral and persistent pools by reducing the adsorption capacity of sediments and

releasing Fe-bound phosphate from sediment sinks.

Given the high sorption capacity of sediments as well as low nutrient content of soils

in the surrounding catchments, I sought to better understand whether nitrogen (N), P

or both were most limiting to within pool productivity and whether responsiveness to

nutrient-additions differed between persistent versus more ephemeral pools. I used 6

h in-situ bottle incubations with a 13

C-enriched NaHCO3 isotopic tracer to measure

rates of charophyte and phytoplankton production in response to nutrient

amendments. I hypothesised that autochthonous production is greater in pools that

become disconnected from groundwater owing to increased concentration of

nutrients as a consequence of evaporation processes compared to pools that remain

connected to alluvial groundwater. Charophyte production was ~2 mg C g-1

DW h-1

regardless of hydrologic status, and order of magnitude greater than phytoplankton

production (~0.01 mg C g-1

DW h-1

). While charophyte productivity was not

significantly increased with either N or P addition, productivity was nevertheless

positively correlated to both charophyte N (R2 = 0.65, p < 0.001) and P tissue content

(R2 = 0.41, p < 0.001). Overall, these findings suggest that P may be more limiting

over longer periods to pool productivity in the Pilbara.

Finally, I investigated if and how periphyton community structure is affected by

increased availability of N and/or P. Given the limited short term responsiveness to

vi

nutrient additions observed in the 13

C-labelling experiment, I paired a 28 day nutrient

limitation experiment using diffusing substrates with photo- and accessory-pigment

analysis to: i) identify which nutrient(s) most limit periphyton production; and ii)

assess if particular components of the periphyton community respond in different

ways to N and P or both. I found that periphyton communities in both persistent and

ephemeral pools were co-limited by N and P availability, which interacted

synergistically. Nitrogen additions caused the periphyton to shift from a diatom-

dominated to chlorophyte-dominated community structure. In contrast, P additions

reduced diatom biomass, and in ‘ephemeral’ pools also promoted dinoflagellate

growth. These findings, together with observations of charophyte and phytoplankton

responses, suggest that autotrophic production is likely co-limited by both N and P.

Community structure within pools, particularly of the periphyton, is also likely to be

vulnerable to shifts in nutrient availability in and around pools, which may arise

seasonally owing to pool contraction during drought or due to impacts of, for

example, livestock activity.

Overall, the findings of the research presented in this thesis demonstrate that

interactions between organic matter and nutrients are important in regulating nutrient

availability in the pools of Pilbara streams. Further, altered nutrient loads and/or

hydrology due to land use change and shifting climate patterns may have significant

but as yet poorly understood impacts on the ecological functioning of intermittent

streams.

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TABLE OF CONTENTS

Thesis Declaration ............................................................................................................. ii

Abstract ............................................................................................................................. iii

Table of Contents ............................................................................................................ vii

List of Tables .................................................................................................................... xi

List of Figures ................................................................................................................. xiii

Acknowledgements ........................................................................................................... xx

Authorship Declaration: Co-Authored Publications ................................................. xxii

1. General Introduction .................................................................................................. 1

1.1. Hydrological characteristics of intermittent and ephemeral streams in hot

arid environments ............................................................................................................. 4

1.2. Processes underpinning nutrient dynamics in intermittent and ephemeral

streams .............................................................................................................................. 8

1.3. Nutrient limitation of metabolic processes in intermittent and ephemeral

streams of hot, arid regions ............................................................................................ 11

1.4. Objectives and organisation of this thesis ........................................................... 13

2. Overview of dryland stream hydrochemistry in the Fortescue River

catchment .......................................................................................................................... 15

2.1. Introduction ......................................................................................................... 15

2.2. Methods ............................................................................................................... 17

2.2.1. Study region and sampling ........................................................................... 17

2.2.2. Stable isotope, carbon, and nutrient analysis .............................................. 24

2.2.3. Data analysis ................................................................................................ 25

2.3. Results ................................................................................................................. 26

2.3.1. Variability in hydrochemistry of surface water across the catchment ......... 26

2.3.2. How is stream nutrient availability linked to hydrology across the

catchment? .................................................................................................................. 34

2.4. Discussion ........................................................................................................... 37

2.4.1. Longitudinal gradients ................................................................................. 37

2.4.2. Evaporative loss drives stream hydrochemistry .......................................... 39

3. Phosphorus sorption and dissolved organic matter interactions in iron-rich

stream sediments .............................................................................................................. 43

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3.1. Introduction ......................................................................................................... 43

3.2. Methods ............................................................................................................... 47

3.2.1. Study site and sampling ............................................................................... 47

3.2.2. Sediment mineralogy and elemental chemistry............................................ 49

3.2.3. Phosphorus sorption characteristics ........................................................... 50

3.2.4. Data analyses ............................................................................................... 53

3.3. Results ................................................................................................................. 54

3.3.1. Sediment properties ..................................................................................... 54

3.3.2. DOM properties of litter leachates .............................................................. 59

3.3.3. Phosphorus sorption characteristics ........................................................... 59

3.3.4. Desorption of P from iron-rich sediments ................................................... 61

3.3.5. Changes in DOM composition with incubation and P adsorption .............. 62

3.4. Discussion ........................................................................................................... 67

3.4.1. DOM composition is influenced by the presence of excess P ...................... 68

3.4.2. Surface/alluvial hydrodynamics do not control sediment P sorption at

the pool scale .............................................................................................................. 69

4. Does low phosphorus limit the short-term metabolic response of

phytoplankton and charophytes of instream pools on an intermittent dryland

stream? .............................................................................................................................. 73

4.1. Introduction ......................................................................................................... 73

4.2. Methods ............................................................................................................... 77

4.2.1. Site description............................................................................................. 77

4.2.2. Pool water physicochemistry of persistent and ephemeral pools ................ 77

4.2.3. Estimation of net ecosystem production ...................................................... 78

4.2.4. Nutrient limitation experiments ................................................................... 78

4.2.5. Laboratory analyses of N, P and carbon ..................................................... 80

4.2.6. Stable isotope analysis of plant tissues, filters and water samples.............. 81

4.2.7. Calculation of productivity based on uptake of 13

CDIC ................................ 82

4.2.8. Data analyses ............................................................................................... 83

4.3. Results ................................................................................................................. 85

4.3.1. Pool hydrology and water chemistry ........................................................... 85

4.3.2. Ecosystem metabolism ................................................................................. 86

4.3.3. 13

C enrichment due to photosynthetic uptake of 13

C-enriched HCO3 .......... 88

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4.3.4. Short-term metabolic response of phytoplankton and charophytes to

nutrient enrichment ..................................................................................................... 89

4.4. Discussion ........................................................................................................... 91

5. Chemotaxonomic responses of autotrophic periphyton communities to

nutrient additions in an intermittent stream ................................................................. 97

5.1. Introduction ......................................................................................................... 97

5.2. Methods ............................................................................................................. 100

5.2.1. Site description ........................................................................................... 100

5.2.2. Nutrient limitation experiments.................................................................. 101

5.2.3. HPLC Pigment analysis ............................................................................. 102

5.2.4. Pool hydrochemistry .................................................................................. 103

5.2.5. Data analyses ............................................................................................. 105

5.3. Results ............................................................................................................... 106

5.3.1. Pool nutrients and hydrologic characteristics ........................................... 106

5.3.2. Periphyton biomass response to nutrient additions ................................... 106

5.3.3. Chemotaxanomic response of autotrophic periphyton .............................. 110

5.4. Discussion ......................................................................................................... 117

6. General discussion .................................................................................................. 120

6.1. Overview ........................................................................................................... 120

6.2. Alluvial groundwater connectivity influences stream biogeochemistry and

metabolism ................................................................................................................... 120

6.3. Sediment mineralogy constrains within-stream nutrient bioavailability ........... 123

6.4. Complex responses of aquatic primary productivity to perturbations in

nutrient status in dryland streams ................................................................................. 126

6.5. Implications from this research to understanding responses of stream

ecosystems in northwest Australia to changing land use and climate ......................... 128

6.6. Conclusion ......................................................................................................... 131

7. Appendix 1 - Pilot study investigating the suitability of 31

P-nmr for the

characterisation of organic phosphorus in iron-rich Pilbara stream sediments ...... 133

7.1. Methods ............................................................................................................. 133

7.1.1. Sample pre-treatment ................................................................................. 133

7.1.2. Sediment chemistry..................................................................................... 134

7.1.3. 31

P-nmr experiment .................................................................................... 134

7.2. Results ............................................................................................................... 136

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7.2.1. Sediment and extract chemistry ................................................................. 136

7.2.2. 31P-nmr spectra......................................................................................... 137

7.2.3. Comments on method suitability ................................................................ 137

8. Appendix 2 – Two-way ANOVA ........................................................................... 139

References ....................................................................................................................... 140

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LIST OF TABLES

Table 2.1 Surface water sampling sites across the Fortescue River catchment,

Northwest Australia. ...................................................................................... 19

Table 3.1 Chemical characteristics of sediments collected from Coondiner Creek.

Electrical conductivity (EC) and pH were measured in a 1:10 (w/v) soil-

solution. Bulk sediment samples from each pool were air-dried prior to

chemical analysis. .......................................................................................... 56

Table 3.2 Freundlich and Langmuir model parameters fitted to experimental

adsorption isotherms. KF: Freundlich adsorption energy constant, n:

Freundlich correction factor, KL: Langmuir isotherm constant (L mg-1

), b:

Langmuir maximum adsorption capacity (mg g-1

). Mean values with standard

deviation in parenthesis (n = 3). Model fits were compared using adjusted-R2,

and residual sum of squares (RSS)................................................................. 58

Table 4.1 Ambient dissolved nitrate/nitrite (NOx), ammonium (NH4), and soluble

reactive phosphorus (SRP) concentrations of stream water at each pool. N:P

ratios calculated as the ratio between DIN and SRP where DIN = NOx + NH4.

Values given are means and standard deviation (n = 3). Ambient dissolved

inorganic carbon (DIC) concentration and its carbon isotope ratio (δ13

CDIC)

are also given.................................................................................................. 79

Table 4.2 C:N:P stoichiometry of pool water, charophytes and phytoplankton. ...... 84

Table 5.1 Characteristics of study pools along Coondiner Creek at initial and final

period of periphyton incubation. Total dissolved nitrogen (TDN), soluble

reactive phosphorus (SRP), dissolved organic carbon (DOC), specific

absorbance at 254nm (SUVA254), dissolved inorganic carbon (DIC), stable

xii

isotopes of filtered water samples (δ13

CDIC, δ2H, and δ

18O), and pool

evaporative loss (f). ...................................................................................... 109

Table 5.2 Peak identification table of pigments identified in mixed standard and

periphyton samples. ..................................................................................... 110

Table 5.3 Factorial two-way mixed effects PERMANOVA of a) periphyton pigment

biomass (µg cm-2

), and b) estimates of algal group contributions from

CHEMTAX analysis of Chl a: Pigment ratios. Pool hydrology and nutrient

treatment are included as factors. Significant P-values are indicated in bold.

...................................................................................................................... 114

Table 7.1 Composition of sediment extracts determined by ICP-OES and

colourimetry. ................................................................................................ 135

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LIST OF FIGURES

Figure 1.1 Pilbara streams show extreme flow intermittency and are typical of those

characterised by Kennard et al. (2010) as ‘variable summer extremely

intermittent’. Daily rainfall (red) and discharge (blue) at river gauging sites in

the a) Upper Fortescue catchment (Fortescue River at Newman 708011), and

b) Lower Fortescue catchment (Fortescue River at Bilanoo pool 708015).

Data source: http://wir.water.wa.gov.au .......................................................... 4

Figure 1.2. Different settings across the catchment include confined gorges with

bedrock substrate at a) Dales Creek, and b) Fortescue River South at

Hamersley Gorge, semi-confined gorges with alluvium at Coondiner Creek c)

and d), spring-fed rivers and streams at e) Fortescue River at Millstream and

f) Weeli Wolli Creek. ....................................................................................... 7

Figure 2.1 a) The Fortescue river catchment of, northwest Australia. b) Location of

sampling sites across the Fortescue River catchment, northwest Australia. c)

Lower and Upper Fortescue River sub-catchments. ...................................... 18

Figure 2.2 a) Stable isotope composition of perennial (red) and intermittent (yellow)

stream waters, located in Lower (□) and Upper (○) catchment positions in the

Fortescue River catchment, northwest Australia. Also shown are rainwater

(+) and groundwater () samples collected from the region at the time of this

study. The local evaporation line (LEL) was calculated from 55 surface water

samples whilst the local meteoric water line (LMWL) was sourced from

Dogramaci et al. (2012). b) Relationship between the distance from river

mouth and stable hydrogen (δ2H) and c) oxygen (δ

18O) isotopes of water. .. 28

xiv

Figure 2.3 Relationship between the distance from river mouth and total dissolved

nitrogen (TDN), soluble reactive phosphorus (SRP), N:P ratio, dissolved

organic carbon (DOC), dissolved inorganic carbon (DIC), and the stable

isotope composition of dissolved inorganic carbon (δ13

CDIC). Shaded fill

denote perennial (red) and intermittent (yellow) stream waters, symbols

denotes Lower (□) and Upper (○) catchment position (n = 55). .................... 31

Figure 2.4 a) Excitation-emission spectra of the three modelled components from

PARAFAC analysis. b) Spectral loadings for excitation (red dash) and

emission (blue line) wavelengths for each component. Components 1 and 2

are humic-like, Component 3 is protein-like tyrosine. c) Relationship

between the stable isotope composition of water (δ18

O) and dissolved organic

matter components. Shaded fill denote perennial (red) and intermittent

(yellow) stream waters, symbols denotes Lower (□) and Upper (○) catchment

position (n = 55). ............................................................................................ 33

Figure 2.5 Principal component analysis of hydrochemical parameters of perennial

(red) and intermittent (yellow) stream waters, symbols denotes Lower (□)

and Upper (○) catchment position (n = 55). The first two axis of the PCA

explains 50.5 % of the total variance. Water stable isotopes δ2H, δ

18O and

δ13

C-DIC contribute consistently to PC1, dissolved organic matter

components C1, C2, and contribute to PC2. .................................................. 34

Figure 2.6 Relationship between the stable isotope composition of water (δ18

O) and

total dissolved nitrogen (TDN), soluble reactive phosphorus (SRP), N:P ratio,

dissolved organic carbon (DOC), dissolved inorganic carbon (DIC), and the

stable isotope composition of dissolved inorganic carbon (δ13

CDIC). Shaded

xv

fill denote perennial (red) and intermittent (yellow) stream waters, symbols

denotes Lower (□) and Upper (○) catchment position (n = 55). .................... 36

Figure 3.1 a) The Fortescue river catchment (solid fill) of the semi-arid Pilbara

region (hatching), northwest Australia. b) Location of Coondiner Creek in the

Upper Fortescue River catchment, c) ‘persistent’ (black squares) and

‘ephemeral’ (grey circles) pools sampled along Coondiner Creek. ............... 48

Figure 3.2 Non-metric multidimensional scaling (nMDS) plots of a) elemental

composition of sediments from XRF, and b) mineralogy of sediments from

XRD for ‘persistent’ (black squares) and ‘ephemeral’ (grey circles) pools of

Coondiner Creek. Data were normalised prior to scaling. ............................. 54

Figure 3.3 Fluorescent DOM components derived from fluorescence spectroscopy

and PARAFAC analysis. a) Modelled excitation-emission spectra of humic-

like components 1 and 2, protein-like component 3, and unknown component

4. b) Excitation (red dash) and emission (blue line) spectral loading of each

corresponding component. ............................................................................. 57

Figure 3.4 Experimental data from batch phosphorus adsorption experiments fitted

to Freundlich (solid line) and Langmuir (dash) isotherms. Mean adsorption

(qe) and standard error (n = 3) for sediments from ‘persistent’ (black square)

and ‘ephemeral’ (grey circle) pools versus equilibrium P concentration (Ce)

shown. ............................................................................................................ 60

Figure 3.5 Phosphorus adsorption (Pads) versus desorption (Pdes) patterns of

Pilbara sediments. Values shown are means with standard error (n = 3) of

sediments from ‘persistent’ (black square) and ‘ephemeral’ (grey circle)

pools. .............................................................................................................. 61

xvi

Figure 3.6 Dissolved organic carbon (DOC), C:N ratio, and Specific UV absorbance

at 254 nm (SUVA254), at the conclusion of batch phosphorus adsorption

experiments. Values shown are means with standard error (n = 3) for


pools. Note different scales on y-axis between DOC panels. ........................ 63

Figure 3.7 UV-vis and fluorescence indices measured at the conclusion of batch

phosphorus adsorption experiments. Spectral slope (S275-295) and humification

index (HIX) values are presented as means with standard error (n = 3) for


pools. .............................................................................................................. 65

Figure 3.8 Fluorescence maxima (Fmax) for PARAFAC derived DOM components at

the end of 24 h batch phosphorus adsorption experiments for sediments from

‘persistent’ (black) and ‘ephemeral’ (grey) pools. X-axis indicates the initial

P concentration (Ci) in batch experiments and y-axis indicates fluorescence

maxima (Fmax) of DOM components. Components 1 and 2 are humic-like

fluorophores, component 3 is protein-like (amino acids), whilst component 4

is thought to be a sediment derived OM degradation product. Values are

given as mean and standard error (n = 3). ...................................................... 66

Figure 4.1 PCA ordination diagram of ‘persistent’ (black squares) and ‘ephemeral’

(grey circles) pools of Coondiner Creek and environmental variables. Cond:

Electrical conductivity , TSS: total suspended solids, TDN: total dissolved

nitrogen, δ2H: water stable isotope deuterium, DIC: dissolved inorganic

carbon, 13C-DIC: δ13

CDIC, DOC: dissolved organic carbon, DO(avg): average

dissolved oxygen, DO(range): dissolved oxygen range, NH4: ammonium,

Temp(avg): average water temperature, Temp(range): Water temperature range,

xvii

NOx: nitrate/nitrite, SRP: soluble reactive phosphorus, Chl a: Chlorophyll a

........................................................................................................................ 85

Figure 4.2 Diel dissolved oxygen curves for stream pools along Coondiner Creek.

Values for gross primary productivity (g O2 m-3

d-1

) (GPP), community

respiration (CR24), net ecosystem production (NEP), and GPP:CR ratio are

given on the figure for each pool. Grey and light shading indicate night and

day periods. Boxed areas signify time envelope for bottle assays. ................ 87

Figure 4.3 Carbon stable isotope ratios of phytoplankton under light/dark conditions

incubated in situ with and without 13C-enriched HCO3 added. Bars are

means and error bars indicate standard error (n = 3). .................................... 89

Figure 4.4 Short-term productivity response of charophytes and phytoplankton to

nutrient additions in ‘persistent’ and ‘ephemeral’ pools estimated as rate of

13C-enriched HCO3 uptake. Bars are means and error bars indicate standard

error (n = 3). ................................................................................................... 90

Figure 4.5 Relationship between the rate of production and tissue content in

charophytes at the end of the incubation experiment. Charophyte content of

a) nitrogen, (%) b) phosphorus, and c) nitrogen:phosphorus (N:P) ratio....... 91

Figure 5.1 Periphyton chlorophyll a response to nutrient additions in ‘persistent’ and

‘ephemeral’ pools. Nutrients added to substrates were nitrogen (N) as

NH4NO3, phosphorus (P) as KH2PO4, and nitrogen + phosphorus (NP). The

control (C) received no nutrient additions. The experiment was duplicated

with ‘grazed’ and ‘ungrazed’ NDS treatments............................................. 108

Figure 5.2 HPLC chromatograms showing a) standard pigment mix, peak numbers

correspond with those in Table 5.2, and b) a typical HPLC chromatogram

from a persistent pool showing control (black), nitrogen (red), phosphorus

xviii

(blue), nitrogen + phosphorus (green). Absorbance was measured at 450 nm

...................................................................................................................... 112

Figure 5.3 Multidimensional dbRDA plots of pigments extracts from the periphyton

NDS experiment; a) pigment biomass (µg cm-2), and b) estimates of algal

group proportions by CHEMTAX analysis. Results are based on a Bray-

Curtis similarity matrix of log(x + 1) transformed samples (n = 60). ......... 114

Figure 5.4 Estimates of algal group contributions to periphyton community structure

calculated from Monte Carlo perturbations of CHEMTAX analysis. Nutrients

added to substrates were nitrogen (N) as NH4NO3, phosphorus (P) as

KH2PO4, and nitrogen + phosphorus (NP). The control (C) received no

nutrient additions. Mean proportion of each group per nutrient and hydrology

treatment is shown with standard error (n = 3). ........................................... 116

Figure 6.1. Examples of the diversity of hydrologies and settings of streams in the

central Pilbara. Weeli Wolli Creek, a spring-fed creek in the Hamersley

Ranges during a) dry periods receiving minewater discharge, and b) moderate

flood after a 30 mm rainfall event. Note the significant increase in suspended

sediments during flood events. c) and d) Typical catchment vegetation in the

Hamersley Ranges. Many hilltops and slopes have sparse vegetation on

highly weathered skeletal soils. ................................................................... 130

Figure 7.1 a) Comparison between total P measured by ICP-OES and inorganic P

measured by colourimetric detection. Dashed line indicates 1:1 relationship,

solid line indicates linear regression (R2 = 0.98), b) Comparison between

total P and Fe, c) P and Ca (note: log10 scale on y-axis), and d) Al and Ca

(note: log10 scale on y-axis) measured by ICP-OES. Treatments T1:

xix

NaOH+EDTA, T2: dithionite before, T3: dithionite before and after, see

methods for detail. ........................................................................................ 136

Figure 7.2 Solution 31

P-nmr spectra of NaOH-EDTA soil extract from Window pool

(WINA-t1), Coondiner creek. Prepared on a) Brucker 500 in a 10 mm tube,

and b) Brucker 600 in a 5 mm tube. The vertical scale has been exaggerated

10x on the upper trace to delineate individual peaks. .................................. 137

xx

ACKNOWLEDGEMENTS

This research was supported by an Australian Government Research Training

Program (RTP) Scholarship at The University of Western Australia and a

RangelandsNRM Pilbara Corridors Biodiversity Scholarship. The School of

Biological Sciences (formerly School of Plant Biology) provided me with a

computer and some travel support during my candidature. Financial contributions to

fieldwork and laboratory expenses were provided by Australian Research Council

Linkage Grant LP120200002 (Grierson et al.) in collaboration with Rio Tinto, and

via funding from Pilbara Corridors (RangelandsNRM). Thank you to the Australian

Freshwater Science Society (formerly ASL) for providing a student travel grant

during my candidature.

My supervisors were Dr Pauline Grierson, Dr Neil Pettit and Prof Peter Davies. I

would especially like to thank Pauline and Neil for their instrumental guidance and

feedback throughout this research journey. I am immensely grateful for your time

over the past five years. Thanks also to Pauline and Neil’s partners, J.T. and Anne,

for accommodating me during visits to Perth and Albany.

A large consortium of scientists from a number of agencies has been involved in

making this research project happen. Thank you to Dr Michael Donn for your

support and providing access to analytical instrumentation at CSIRO Land and

Water (Chapters 2 & 3). Thanks also to Dr Grzegorz Skrzypek for advice on water

isotopes and understanding evaporative loss. Ian Cotton, Theresa Belcher, and Bill

Crotching organised a number of regional workshops with RangelandsNRM.

Shawan Dogramaci, Sam Luccitti, and Naoko Zwingmann from Rio Tinto Iron Ore

gave logistical support along with XRD/XRF analysis of sediment samples in

Chapter 3. Neil Brougham and Dan Petersen of the WA Department of Parks and

Wildlife accommodated us and gave regional advice whilst in the field. Kate Bowler,

Doug Ford, and Ela Skrzypek of the West Australian Biogeochemistry Centre at

UWA provided stable isotope and nutrient analysis. Greg Cawthray (UWA) gave

critical technical advice with HPLC analysis for Chapter 5. Sara Lock, Samantha

xxi

Lostrom, Dr Jennifer Kelley, Dr Renee Gruber, Doug Ford, Dr Andre Siebers, Neil,

and Pauline gave assistance in the field.

Thank you to current and past members of the Ecosystems Research Group at UWA

who have lent a hand, given advice, been a soundboard, and provided a fun and

nurturing work environment: Dr Alison O’Donnell, Dr Gerald Page, Dr Alex

Rouillard, Dr Andre Siebers, Dr Tegan Davies, Dr Rachel Argus, Doug Ford,

Belinda Martin, Caroline Mather, Hannah Etchells, Jen Middleton, and Josh Oliver.

Thank you to the TropWater group at JCU who welcomed me and provided office

space and IT support during my final write-up. Special thanks to my trusty bicycle

for insistently reminding me no matter how long, steep or daunting, a hill is always

worth climbing at the very least for the perspective from the top, and the thrill of the

descent down the other side.

This project would not have been completed without the support and love of the

people close to me who I am thankful to call family and friends. My Iles and Gruber

families have been a constant source of motivation. Special shout-out to Micha

Campbell, Dr Lies Notebaert, Dr Stijn Masschelein, Dr Patrick Clarke, Dr James

Hitchcock, Nina Gallo, Mark Hamilton, Sam Vinton-Boot, and Freo bicycle riding

buddies: especially Zoe, Paul, Amy, Nick, Tim, Nate, Kiera, and Heath.

And finally, Renee, my love and my partner in all adventures great and small. Your

encouragement and unwavering belief in me was instrumental in completing this

task. Thank you for sharing the bumps and supporting me through this research

journey. So, what's next?

xxii

AUTHORSHIP DECLARATION: CO-AUTHORED PUBLICATIONS

This thesis contains work that has been prepared for publication. This thesis does not

contain work that I have previously published, nor work under review for publication

at the time of thesis submission.

Details of the work:

Iles, J.A., Skrzypek, G., Pettit, N.E., Grierson, P.F. (in prep) Hydrochemistry of

dryland streams in an arid-zone catchment

Location in thesis:

Chapter 2 – Overview of dryland stream hydrochemistry in the Fortescue River

catchment

Student contribution to work:

JAI contributed to study design, undertook all experimental work, analysed the data,

and wrote the manuscript. GS contributed to isotope analysis and interpreting results.

NEP contributed to study design and assisted with interpreting results. PFG

contributed to study design and assisted with interpreting results. PFG and NEP

commented on the manuscript.

xxiii


Iles, J.A., Donn, M.J., Pettit, N.E., Grierson, P.F. (in prep) Phosphorus sorption and

dissolved organic matter interactions in iron-rich stream sediments

Location in thesis:

Chapter 3 – Phosphorus sorption and dissolved organic matter interactions in iron-

rich stream sediments



and wrote the manuscript. MJD assisted with sample analysis and interpreting

results. NEP contributed to study design and assisted with interpreting results. PFG

helped conceptualise the study and contributed to study design and assisted with

interpreting results. PFG and NEP commented on the manuscript.


Iles, J.A., Pettit, N.E., Grierson, P.F. (in prep) Does low phosphorus limit the short-

term metabolic response of phytoplankton and charophytes in an intermittent dryland

stream?

Location in thesis:

Chapter 4 – Does low phosphorus limit the short-term metabolic response of

phytoplankton and charophytes in an intermittent dryland stream?



and wrote the manuscript. NEP and PFG contributed to study design and assisted

with interpreting results. PFG and NEP commented on the manuscript.

xxiv


Iles, J.A., Cawthray, G.R., Pettit, N.E., Grierson, P.F. (in prep) Chemotaxonomic

responses of autotrophic periphyton communities to nutrient additions in an

oligotrophic intermittent stream

Location in thesis:

Chapter 5 – Chemotaxonomic responses of autotrophic periphyton communities to

nutrient additions in an oligotrophic intermittent stream



and wrote the manuscript. GRC assisted with HPLC experiment design and analysis.

NEP and PFG contributed to study design and assisted with interpreting results. PFG

and NEP commented on the manuscript.

Student signature:

Date: 31th

August 2018

I, Pauline Grierson certify that the student statements regarding their contribution to

each of the works listed above are correct

Coordinating supervisor signature:

Date: 31 August 2018

1

1. GENERAL INTRODUCTION

Intermittent and ephemeral streams dissect the arid and semi-arid landscapes

(drylands) of the world, providing unique habitat niches in what is frequently an

otherwise hostile landscape. Temporal and spatial heterogeneity in the hydrology of

dryland streams can produce higher ecological diversity than might be expected for

such regions due to increased ecological niche and species turnover, which breaks

down boundaries between lotic, lentic, and terrestrial phases in these streams (Datry

et al., 2014; Acuña et al., 2015; Leigh & Datry, 2017). The remote Pilbara of

northwest Australia is one region known to support high biological richness and

species endemism (Pepper et al., 2013), including of aquatic organisms (Morgan &

Gill, 2004; Reeves et al., 2007; Pinder et al., 2010). The Pilbara is also rich in

mineral resources, and provides an estimated 39 % of the world's iron ore, much of

which is mined below water table (DJTSI, 2018). Consequently, freshwater

ecosystems in the Pilbara are under increasing pressure from resource extraction, as

well as municipal water extraction to support regional development (DOW, 2010;

EPA, 2014). While freshwater ecosystems across the Pilbara have long been subject

to disturbance associated with pastoral activities (see van Vreeswyk et al., 2004;

Halse et al., 2007 for overviews), impacts have generally been localised, and mainly

associated with the direct grazing of vegetation, physical trampling and inputs of

nutrients from manures (Masini, 1988). In contrast, shifts in the timing and volumes

of water flows as well as physical changes in stream morphology resulting from

resource extraction – particularly from mine dewatering – have significantly altered

stream hydrology (Gardiner, 2003; Barber & Jackson, 2011, 2012; Dogramaci et al.,

2

2015). These changes in hydrology can be both highly localised as well as

cumulative in their impacts on freshwater ecosystems across catchments

(Voeroesmarty et al., 2010; EPA, 2014). Consequently, a rigorous understanding of

how such these ecosystems function is critical in order to better predict the risks of

anthropogenic changes to the ecology of intermittent and ephemeral streams.

As for many arid zones around the world, freshwater ecosystems of the Pilbara are

both spatially and temporally highly variable (Kennard et al., 2010), and stream

biogeochemical processes are strongly mediated by an episodic hydrology (Siebers

et al., 2016). The soils that dominate the catchments across the Pilbara, where the

research described in this thesis is focused, have developed from some of the oldest

erosion surfaces on Earth, including iron-rich sedimentary deposits (2.77 - 2.4 Ga)

that have been uplifted to form the Hamersley Range (Kranendonk et al., 2002;

Arndt et al., 2007). Consequently, Pilbara soils tend to be heavily weathered, are

frequently dominated by smectite clays, and contain relatively low levels of available

nutrients and organic matter (Bentley et al., 1999; Islam & Adams, 2001; McIntyre

et al., 2009a; McIntyre et al., 2009b). Aquatic productivity of streams across the

region is therefore likely to be strongly limited by nutrient supply, particularly of

nitrogen (N) and phosphorus (P).

In largely undisturbed dryland catchments, cycles of flood and drought control fluxes

of nutrients and organic matter into streams and rivers (Bunn et al., 2006b; Leigh et

al., 2010). However, this natural biogeochemical variability remains largely

unquantified in northwest Australia and thus disentangling the impacts of altered

land use from background variability is challenging. Nevertheless, understanding of

3

both regional and more localised hydrology in the Pilbara has been improved by the

study of hydrogeochemical processes (Dogramaci et al., 2012; Skrzypek et al., 2013;

Mather et al., 2018) as well as reconstruction of past climates from tree rings

(O'Donnell et al., 2015) and flood regimes from sediments and other records

(Rouillard et al., 2015; Rouillard et al., 2016). Recent studies by Fellman et al.

(2011) and Siebers et al. (2016) have highlighted the strong influence of

hydrological connectivity to groundwater on dissolved organic matter (DOM)

biogeochemistry of stream pools in the Pilbara, with consequent impacts on trophic

structure (Siebers, 2015). However, interactions between nutrient cycling processes

and productivity of the different pools and reaches of dryland streams remain poorly

understood, not only in the Pilbara but also across many other arid regions of the

world (Mulholland & Webster, 2010; Bernhardt et al., 2018).

In this thesis, I have used field surveys coupled with manipulative experimental

approaches in both the field and laboratory to further a mechanistic understanding of

nutrient and carbon biogeochemistry of dryland streams of the Pilbara. This work is

also applied, and will help elucidate the role of hydrologic connectivity in

maintaining ecosystem function, a key concern for future management of these

systems. This introductory chapter provides a general overview of the hydrological

characteristics of dryland streams and rivers, briefly outlines the current

understanding of nutrient uptake and limitation in these streams, and provides the

context for the following experimental chapters.

4

1.1. Hydrological characteristics of intermittent and ephemeral streams in

hot arid environments

Intermittent rivers and ephemeral streams (IRES) are streams that only flow during

the periods when they receive water from springs or surface runoff, and cease

flowing during dry periods, often seasonally (Gordon et al., 2004). Locally elevated

water tables or perched aquifers may also maintain surface water during dry periods

in some intermittent streams. In contrast, an ephemeral stream only flows in direct

Figure 1.1 Pilbara streams show extreme flow intermittency and are typical of those

characterised by Kennard et al. (2010) as ‘variable summer extremely intermittent’. Daily

rainfall (red) and discharge (blue) at river gauging sites in the a) Upper Fortescue catchment

(Fortescue River at Newman 708011), and b) Lower Fortescue catchment (Fortescue River

at Bilanoo pool 708015). Data source: http://wir.water.wa.gov.au

5

response to rainfall events, such as from thunderstorms, and are typically dry for

most of the year. Intermittent and ephemeral streams make up > 50 % of

watercourses globally (Datry et al., 2016), and more than 65 % of streams of

continental Australia (De Vries et al., 2015; Datry et al., 2018a), although this is

likely an underestimation as intermittent streams are generally poorly represented in

stream gauging networks (Acuña et al., 2014). A desktop mapping approach

estimated 82 % of major streams in Australia (by stream length) as intermittent

(Geofabric, 2012), although this value increases to 98 % if also including all minor

headwater and more ephemeral streams. In light of their widespread geophysical

significance, IRES are globally underrepresented in scientific studies (Stubbington et

al., 2018).

Much of our understanding of the functioning of dryland streams within Australia

has developed from studies of lowland rivers of the highly regulated Murray-Darling

Basin (Kingsford, 2000; Mitrovic et al., 2003) and from the vast braided, flat

landscapes of inland central Australia, such as Cooper Creek (Bunn et al., 2003;

Fellows et al., 2007). However, the streams of the inland Pilbara are characterised as

extremely intermittent under a climate of highly variable rainfall (Kennard et al.,

2010). Surface water is driven by runoff associated with high intensity rainfall events

resulting from cyclonic activity and tropical lows through the austral summer

(Ruprecht, 1996). While mean annual discharge at the mouth of the Fortescue River,

one of the largest catchments in the region, is 292 GL y-1

flows are extremely

variable (range: 0 to 1420 GL y-1

, http://wir.water.wa.gov.au) (Figure 1.1). Gauging

sites across the Fortescue catchment are generally located at more persistent sites of

surface water such as large river pools, whilst intermittent streams are poorly

6

represented across the network. Where hydrological gauging data are available, a

rapid rising and falling limb of the hydrograph is observed following significant

rainfall events, hence surface flows are often intense but short lived

(http://wir.water.wa.gov.au). However, it is these larger precipitation events (> 20

mm day-1

) that drive groundwater recharge, primarily through the alluvium

(Dogramaci et al., 2012). These groundwater reserves then maintain surface water

within streams as the region progressively dries and streams contract to a series of

pools, such that the degree of alluvial connectivity is an important characteristic

which governs biogeochemical processes within these pools (Fellman et al., 2011;

Siebers et al., 2016). Figure 1.2 illustrates some examples of stream types and

landscapes found in the Pilbara.

7

Figure 1.2. Different settings across the catchment include confined gorges with bedrock

substrate at a) Dales Creek, and b) Fortescue River South at Hamersley Gorge, semi-

confined gorges with alluvium at Coondiner Creek c) and d), spring-fed rivers and streams at

e) Fortescue River at Millstream and f) Weeli Wolli Creek.

8

1.2. Processes underpinning nutrient dynamics in intermittent and

ephemeral streams

Instream processes within IRES are different to perennial systems in a number of

ways. Biophysical and ecological features of IRES are largely determined by

patterns of flow intermittence (Datry et al., 2017). Dying and rewetting cycles

resulting from the highly dynamic flows described above can be especially important

in shaping sediment geochemistry (Baldwin & Mitchell, 2000) and terrestrial organic

matter (Baldwin, 1999; Datry et al., 2018b), as well as stimulating biodiversity

(Leigh & Datry, 2017). Rewetting cycles are especially important in delivering

nutrients and carbon from the surrounding catchment, and not only in hot, arid

catchments. For example, high flows during large winter storms reconnect dried

reaches and stimulate decomposition in temperate streams (Northington & Webster,

2017). In contrast, drying and UV photo-degradation of organic matter are much

stronger influences on decomposition rates in hot dryland regions (Fellman et al.,

2013), especially when iron-oxides are present (Howitt et al., 2008). Evapo-

concentration of nutrients and carbon is also a significant process on surface waters

in hot arid regions, especially once streams cease to flow and fragmentation occurs

(Sheldon & Fellows, 2010; Siebers et al., 2016). Stream sediments regulate P content

of the water column through sorption processes and thus has a strong influence on

nutrient cycling in aquatic systems (Reddy et al., 1999). Transport and subsequent

reworking of sediments during flood flows produces ‘fresh’ material for P sorption,

whilst potentially transporting sediment-P from the active hyporheic zone and

depositing elsewhere.

9

Downstream transport of nutrients and carbon is restricted to short periods of flow in

Pilbara streams. Once flow ceases and stream pools become progressively

disconnected, allochthonous sources of energy become exhausted and the internal

cycling of nutrients and carbon becomes increasingly important. Autochthonous

production by macrophytes, phytoplankton, and periphyton increasingly maintains

consumer food webs during the longer no-flow period. However, current nutrient

transport models are inadequate for describing longitudinal nutrient dynamics within

these dryland streams. For example, the concept of nutrient spiralling, whereby

nutrients are cycled as they progress downstream (Webster, 1975; Newbold et al.,

1982) is not applicable when downstream transport occurs via intermittent flow

(Fisher et al., 2004). Previous attempts to refine the nutrient spiralling concept to

include flow variability have been unsuccessful at incorporating the extreme end of

the flow variability spectrum seen in ephemeral systems (Fisher et al., 1998).

Consequently, further exploration of river and stream functioning in tropical and

sub-tropical regions especially, should contribute to expanding such models to

include more extreme systems.

In intermittent streams of hot and arid regions, sediments are frequently dry for

periods from months to years, such that terrestrial plant litter accumulates both

within pools along the stream bed and in the adjacent parafluvial zones (Datry et al.,

2018b). When dry sediments are rewetted, organic matter is mineralised and there is

a release of nutrients and carbon into pore water and the water column (Baldwin &

Mitchell, 2000). In the Pilbara, these initial ‘flashy’ flows are highly important to the

productivity of aquatic ecosystems as they provide an opportunity for nutrients and

carbon from the surrounding catchment to enter the stream. In contrast, streams in

10

more mesic environments receive a more consistent input of materials throughout the

year via baseflow (Buffam et al., 2001; Bieroza & Heathwaite, 2016). Consequently,

productivity in Pilbara streams and other IRES in Australia is thought to be primarily

driven by allochthonous inputs of nutrients and organic matter sourced from the

catchment in the form of organic matter derived from vegetation and soils (Siebers,

2015). In contrast, food webs which lack significant inputs of allochthonous material

may revert to autochthonous algal production (Bunn et al., 2003).

Continental Australian soils are highly weathered and nutrient poor (Orians &

Milewski, 2007), especially in phosphorus (P) (Holford, 1997; Doolette et al., 2011).

Source rock and soils of the Pilbara region are especially depauperate in P compared

to many parts of the world, and available nitrogen (N) is also frequently low (Ford et

al., 2007; McIntyre et al., 2009b). Inputs of N via biological fixation processes are

limited by low P as well as aridity. Consequently, rates of terrestrial primary

production in dry arid regions are also low. Hence, the quantity of nutrient and

carbon available for transportation to streams from surrounding catchments is

limited. Therefore, bioavailable forms of N and P in surface waters across the Pilbara

are also generally very low (Pinder et al., 2010; Fellman et al., 2011) compared to

other Australian streams (Harris, 2001), and to streams globally (Smith et al., 2003;

Seitzinger et al., 2010).

Streams across the Pilbara region have a lower nutrient status compared to studies

elsewhere that have examined biogeochemical processes in IRES such as Europe

(Acuna et al., 2004; von Schiller et al., 2011) and North America (Grimm, 1992;

Sponseller & Fisher, 2006); they therefore offer novel conditions to explore nutrient

11

processes in IRES and extend knowledge of how ecosystems may be adapted to such

limitation. For example, how do stream metabolic processes occur in oligotrophic

(low nutrient) systems, especially under a highly intermittent hydrology? What

biogeochemical processes determine the availability of N and P at any one time, and

conversely what processes are most limited under these oligotrophic conditions?

How do they contrast with other IRES world-wide?

1.3. Nutrient limitation of metabolic processes in intermittent and

ephemeral streams of hot, arid regions

Ecosystem metabolism comprises the processes of productivity and respiration in

aquatic ecosystems (Odum, 1956; Staehr et al., 2011). Photosynthesis drives carbon

fixation and productivity during daylight hours, whilst both autotrophs and

heterotrophs respire. The degree of autotrophy or heterotrophy (ratio of gross

production to respiration) of a system has been interpreted elsewhere as an indicator

of overall ecosystem health (Fellows et al., 2006; Likens et al., 2009). Whilst light

and carbon inputs are important sources of energy of instream metabolism, the

ability of aquatic primary producers to acquire inorganic nutrients is a major control

on production in all systems.

The rate of primary production is controlled by Liebig’s law of the minimum (de

Baar, 1994) where the most limiting nutrient or energy source in a system determines

this rate. In freshwater systems, we are primarily interested in limitation by nitrogen

and/or phosphorus (Francoeur, 2001; Tank & Dodds, 2003; Elser et al., 2007). Much

of the pioneering research on biogeochemical processes in intermittent streams

focused on nitrogen limited arid systems, such as Sycamore Creek in Arizona

12

(Grimm et al., 1981; Grimm & Fisher, 1986). These streams are fed by seasonal

snow-melt, are frequently flow regulated by lock or flood controls, and therefore

have a comparatively predictable hydrology. Less is understood of the relative

importance of nitrogen versus phosphorus availability in IRES that occur in largely

unmodified catchments, receive very episodic flows, and where the surrounding

catchments are dominated by nutrient-poor soils. Factors controlling stream

productivity in intermittent Pilbara streams remain largely undescribed. However,

expanding our understanding of metabolic processes in IRES is considered

fundamentally important for the future management of freshwater systems given

there is a global trend of increasing stream intermittency (Acuña et al., 2017). The

Pilbara region in many ways exemplifies these global processes, with changing

spatial and temporal patterns in rainfall, and changed land use (Cullen & Grierson,

2007; O'Donnell et al., 2015; Rouillard et al., 2015; Rouillard et al., 2016).

Nutrient limitation studies have broadened our understanding of water column and

benthic autotrophic production (Francoeur, 2001) and heterotrophic respiration

(Burrows et al., 2015). Nutrient limitation has also been investigated utilising a range

of approaches including whole lake fertilisation studies (Carpenter et al., 2001),

mesocosm experiments (O'Brien & Dodds, 2007), and incubations of in situ nutrient

diffusing substrates (Fairchild et al., 1985; Tank & Dodds, 2003; Capps et al., 2011).

Typically changes in biomass or chlorophyll-a are measured to assess how primary

producer growth may respond to nutrient addition (thus indicating limitation).

However, if multiple algal species are present they may not respond uniformly to

nutrient addition (e.g. N versus P). For example, freshwater cyanobacteria containing

heterocysts have the ability to fixate atmospheric nitrogen (N2) under nitrogen

13

starvation (Carey et al., 2012). Nitrogen-fixers would be expected to show little

response to N additions though may boom under elevated P (Cottingham et al.,

2015). Consequently, studies that also examine shifts in community composition and

abundance rather than purely total periphyton production may provide greater insight

into overall nutrient limitation in any one system or time (Townsend et al., 2012;

Dalton et al., 2015).

1.4. Objectives and organisation of this thesis

The general objective of this thesis is to increase understanding of how the

intermittent and ephemeral streams of hot arid environments function ecologically,

and especially how phosphorus and nitrogen may interact to influence aquatic

metabolism in the Pilbara region of northwest Australia. I examined patterns and

processes in stream biogeochemistry in the context of stream hydrology at the

catchment to reach scale. Specifically, I sought to: i) describe the biogeochemical

characteristics of dryland streams in the Fortescue River catchment as context for

understanding the broad-scale variability in nitrogen and phosphorus availability in

relation to groundwater connectivity; ii) investigate key geochemical processes that

are likely to influence nutrient transfers between the iron and calcium-rich sediments

and the water column; iii) quantify aquatic primary production response to N and P

additions; and iv) consider the influence that connectivity to water in the alluvium

has on all of the above processes.

The research incorporates both field observations and experimental results made

over a three year period from May 2013 to October 2016. The thesis is presented as a

14

series of ‘stand-alone’ journal papers; as such, some repetition is unavoidable. I have

truncated methods and referred to their full explanation in prior chapters where it

does not detract from the flow of the text. References to in-text citations for each

chapter are compiled at the end of the thesis. Chapter 2 presents a study of the

variation in stream geochemistry across a large arid-zone catchment, and provides a

regional context for the following experimental chapters, which are then focused on

Coondiner Creek in the Upper Fortescue River sub-catchment. Chapter 3

investigates adsorption kinetics influencing phosphorus (P) bioavailability and

emphasises interactions between P, dissolved organic matter (DOM) and iron (Fe).

Chapter 4 presents a study on the short-term metabolic response of aquatic primary

producers to nutrient additions using 13

C isotopic labelling approaches. Chapter 5

investigates the response of periphyton communities to nutrient additions using

nutrient diffusing substrate and subsequent pigment analysis. Finally, Chapter 6

provides a general discussion of the overall work in the context of the functioning

and importance of intermittent and ephemeral streams both in Australia and

elsewhere, the implications of this study for assessing how changing climate,

hydrology, and land use can alter the biogeochemical character of these streams, and

opportunities to mitigate and minimise future anthropogenic changes.

15

2. OVERVIEW OF DRYLAND STREAM HYDROCHEMISTRY IN

THE FORTESCUE RIVER CATCHMENT

2.1. Introduction

Intermittent rivers and ephemeral streams (IRES), which periodically cease to flow,

comprise around 50% of the global fluvial network (Datry et al., 2017), yet are

under-represented in stream monitoring; even basic information on surface flows and

nutrient levels is generally lacking (Acuña et al., 2014). While IRES are known for

their extreme temporal hydrologic variability, hydrologic regimes in dryland

catchments are also frequently spatially heterogeneous. Dryland catchments are a

mosaic of not only intermittent and ephemeral reaches but may also contain

perennial reaches, and isolated pools maintained by alluvial groundwater.

Consequently, dryland catchments likely contain a continuum of hyporheic flow

paths that are associated with different hydrologic residence times (Boano et al.,

2014). These hyporheic zones, where surface and ground waters are exchanged, are

important for vertical and lateral connections of rivers (Ward, 1989; Thorp et al.,

2006), and play a significant role in biogeochemical cycling of carbon and nutrients

(Boulton et al., 1998; Fellman et al., 2011; Siebers et al., 2016). However, most

studies of hyporheic processes in IRES have focused on relatively small scales;

understanding of hyporheic exchange at larger scales e.g. across a catchment, is

largely lacking (Magliozzi et al., 2017).

The Pilbara region of northwest Australia typifies the challenge of acquiring better

representation of IRES-dominated catchments - the region is sparsely populated,

16

remote, largely hydrologically unregulated, and poorly represented in hydrographic

monitoring networks. Groundwater is also recognised as playing a key role in

maintaining perennial and persistent pools and reaches across the catchment

(Baimbridge et al., 2010; Siebers et al., 2016).

Where detailed monitoring data may be lacking, stable isotopes of water (δ2H and

δ18

O) have proved particularly useful tools for understanding key hydrologic

processes and identifying zones where hyporheic exchange is particularly important.

For example, the isotopic compositions of stream waters in arid regions, including

the Pilbara, has been shown to be strongly determined by fractionation due to

evaporative pressure (Dogramaci et al., 2012), as well as the degree of connectivity

to alluvial water (Fellman et al., 2011). When water isotope data are coupled with

other measures of hydrochemistry, it should thus be possible to assess if any patterns

are evident at the catchment scale.

In this chapter, I provide an overview of the hydrochemical characteristics of the

Fortescue River catchment, the largest catchment of the central Pilbara region. This

overview is primarily intended to provide broad scale context for subsequent

experimental chapters (3-5), which are focussed on understanding how nitrogen and

phosphorus availability/limitation, as well as dissolved organic matter dynamics, are

influenced by connectivity to groundwater at more localised scales. I was

particularly interested in determining if longitudinal gradients (upstream to

downstream) are evident in the Pilbara as observed elsewhere in catchments with

more perennial flows (Vannote et al., 1980). Specifically, I sought to discover if

there is any evidence of longitudinal gradients in a) stream water residence time, b)

17

nutrient concentrations and c) form and concentrations of dissolved carbon across

the catchment.

2.2. Methods

2.2.1. Study region and sampling

The Fortescue River is in the Pilbara region of northwest Australia and is 760 km in

length (Figure 2.1). The total catchment is 48,360 km2 in size but is usually

considered in two parts; The upper eastern region of the catchment (Upper Fortescue

River catchment, 29,752 km2) is endorheic draining the gorges and northern flanks

of the Hamersley Ranges in to the Fortescue Marsh (Barnett & Commander, 1985).

The Upper catchment is physiographically separated from the Lower Fortescue River

catchment (18,608 km2) by the Goodiadarrie Hills (> 410 m a.s.l;

http://www.water.wa.gov.au). The two sub-catchments are also considered to be

hydrologically disconnected, with the Fortescue Marsh acting as a terminal wetland

for drainage from the Upper catchment (Skrzypek et al., 2013).

18

Figure 2.1 a) The Fortescue river catchment of, northwest Australia. b) Location of

sampling sites across the Fortescue River catchment, northwest Australia. c) Lower and

Upper Fortescue River sub-catchments.

19

Tab

le 2

.1 S

urf

ace

wat

er s

amp

lin

g s

ites

acr

oss

the

Fort

escu

e R

iver

cat

chm

ent,

Nort

hw

est

Aust

rali

a.

22

The climate across the Fortescue catchment is sub-tropical semi-arid. Rainfall occurs

predominantly in the austral summer arising from cyclones, monsoonal lows and

tropical thunderstorms, which punctuate periods of prolonged drought (Bureau of

Meteorology, 2018). Mean annual rainfall is ~300 mm across the catchment but

highly variable both within and among years (Chapter 1: Figure 1.1). Temperatures

range from mean daily minima and maxima of 25 to 39 °C in the summer, and from

8 to 22 °C in the winter, such that mean annual pan evaporation ranges from 1200 to

2000 mm across the catchment and far exceeds mean annual rainfall (Charles et al.,

2015).

The flow regime in the Fortescue River and its tributaries is directly linked to rainfall

dynamics, with seasonal discharge during the wet summer months (January to

March) and flow only occurring following cyclonic rainfall or large low pressure

rainfall events (Rouillard et al., 2015). Due to these hydrological constraints, the

streams in this region are extremely intermittent (Kennard et al., 2010). During the

drier winter months and years with no cyclone activity, surface waterways become

disconnected through evaporation to form a chain of pools along drainage lines

(Beesley & Prince, 2010; Fellman et al., 2011; Siebers et al., 2016). Groundwater

thus plays an important role in maintaining surface water volume in many of these

stream pools throughout the catchment (Dogramaci et al., 2012), and maintains

perennial flow to some reaches (Baimbridge et al., 2010). Hence, whilst the system

is highly intermittent at the catchment scale, surface water expression at individual

stream reaches may be on a spectrum from perennial to intermittent.

23

A total of 55 surface water samples were collected from streams across the Fortescue

River catchment between May 2013 and April 2014 encompassing the two major

sub-catchments of the Fortescue River (Figure 2.1). Sampled reaches were classified

as ‘perennial’ if they were known to have continuously flowed for at least the

previous ten years, while all other sites were classed as ‘intermittent’ on the basis

that they were not flowing at time of sampling. Triplicate water samples were taken

at each site for analysis of nutrients and carbon, and a single sample per site was

collected for water isotope analysis. Three rainwater samples collected at Millstream

NP and a single groundwater sample from a flowing bore at Coondiner Creek were

collected in April 2014. A summary of pool size and other parameters are

summarised in Table 2.1.

Stable isotope composition of water (δ2H and δ

18O) was measured to assess both the

source and degree of evaporation of water within the catchment (Skrzypek et al.,

2015). Water samples for δ2H and δ

18O isotope analysis were collected in a glass vial

ensuring all headspace was removed. As regional groundwater can be bicarbonate

rich (Dogramaci & Skrzypek, 2015), the concentration of dissolved inorganic carbon

(DIC) and its stable carbon isotope composition (δ13

CDIC) were measured to also

assess relative contributions of groundwater among sites. Water samples for DIC and

δ13

CDIC analysis were field filtered through a sterile 0.2 µm filter (Sartorius Minisart)

into a glass vial, ensuring all headspace was removed. Dissolved organic carbon

(DOC), dissolved organic matter (DOM) and nutrient samples were field filtered

through a 0.45 µm syringe filter (Sartorius Minisart). Samples for total dissolved

nitrogen (TDN) and soluble reactive phosphorus (SRP) were also collected as an

indicator of trophic state and nutrient limitation. Water samples for isotope, nutrient

24

and carbon analysis were immediately refrigerated (4 °C) in the field up until time of

analysis.

2.2.2. Stable isotope, carbon, and nutrient analysis

Stable isotopes of water (δ2H and δ

18O) were measured on a Picarro L1102-i isotopic

liquid water and continuous water vapour analyser (Picarro, Santa Clara, CA,

U.S.A.). All δ2H and δ

18O values are given in per mil [‰ VSMOW] according to

delta notation (Coplen, 1996). Detailed instrument procedures and standard

verification are outlined in Skrzypek and Ford (2014). Carbon isotope of dissolved

inorganic carbon (δ13

CDIC) was measured on a Thermo Delta XL IRMS with

Gasbench II (Thermo Fisher Scientific, Waltham, MA, U.S.A.). All δ13

CDIC values

are given in per mil [‰ VPDB] according to delta notation (Coplen, 1996).

Analysis of DOC and TDN samples was conducted using a Shimadzu TOC-V

analyser coupled with a TNM-1 total nitrogen module (Shimadzu Corp., Kyoto,

Japan). The concentration of DOC and DOM was characterised with absorbance and

fluorescence spectroscopy and parallel factor analysis (PARAFAC) (Stedmon et al.,

2003). Dissolved aromatic carbon content (SUVA254) was calculated as the

absorbance at 254 nm measured on a Shimadzu UV-VIS spectrophotometer divided

by DOC concentration and is reported in the units L mg-1

m-1

(Weishaar et al., 2003).

DOM fluorescence was measured on a Varian Cary Eclipse spectrofluorometer

(Varian Medical Systems, Inc. California USA). An excitation emission matrix

(EEM) was produced for excitation wavelengths 240 to 450 nm at 5 nm intervals

with emission intensities captured from 300 to 600 nm at 2 nm intervals.

25

Concentration of SRP was measured with the modified ascorbic acid method

(Murphy & Riley, 1962; Kuo, 1996).

2.2.3. Data analysis

Stream water, rainwater and groundwater stable isotope samples were plotted on δ2H

and δ18

O biplots. A linear model was fitted to all surface water samples to calculate a

local evaporation line (LEL) for the Fortescue River catchment. The local meteoric

water line (LMWL) developed for the region by Dogramaci et al. (2012) was added

for comparison. I utilised δ18

O values as a proxy for evaporative loss in stream pools

(Skrzypek et al., 2015). Hydrochemical variables (nutrient and carbon data) were

plotted against δ18

O, and linear models fitted. Models were first fitted to the whole

catchment dataset. Next, the dataset was split into perennial and intermittent sites

and modelled separately to investigate if they differed. Overall differences in the

hydrochemistry of perennial and intermittent reaches were assessed using analysis of

variance (ANOVA); variables were tested for normality (Shapiro-Wilk test) and

homogeneity of variance (Bartlett’s test). Variables were log-transformed where

required to meet assumptions.

Parallel factor analysis (PARAFAC) was performed in MATLAB (R2012a) using

the n-way and drEEM (v4.0) toolboxes (Murphy et al., 2013). Raw EEM’s were

corrected for Raman scatter with a MilliQ blank subtraction. Rayleigh peak regions

were also removed prior to modelling. The PARAFAC model was trained to identify

best fit between three to six fluorophore components with the experimental data. The

final three-component model was validated using split-half analysis (Stedmon &

Bro, 2008), and the modelled components compared to previously identified

26

fluorophore components with OpenChrom (v1.3.0 Dalton) and the OpenFluor

spectral library (Murphy et al., 2014). Peaks are presented as maximum fluorescence

intensity (Fmax) values in Raman units for each component.

Principal component analysis (PCA) was performed with R software (R Core Team,

2017) to reveal patterns in the hydrochemical dataset across the catchment. I

conducted a log-transformation on variables that showed skewness. DIC

concentrations were excluded from the PCA as the dataset contained missing values

of this variable. Data were normalised (scaled) and a correlation matrix produced.

2.3. Results

2.3.1. Variability in hydrochemistry of surface water across the catchment

The three rainfall samples collected at Millstream in April 2014 coincided with the

previously established LMWL for the region (Figure 2.2a). Similarly, the

groundwater sample from Coondiner bore had a stable isotope composition of -

60.01/ -9.13 ‰ for δ2H/δ

18O, which is consistent with the regional groundwater

signature (Dogramaci et al., 2012). The isotopic composition of surface water

sampled across the catchment ranged from -59.49 to +34.35 ‰ for δ2H and -9.01 to

+12.01 ‰ for δ18

O (Figure 2.2a). All stream samples plotted below (to the right) of

the LMWL (Figure 2.2a), indicating evaporation. I first used a linear model to

calculate the local evaporation line from stream water samples. The initial stable

isotopic composition of stream water prior to evaporation was estimated from where

the LEL intersected with the LMWL: -58.7 ‰/9.03 ‰ for δ2H /δ

18O, closely

matching the groundwater signature reported above. Water from intermittent reaches

27

was significantly more enriched in both 2H and

18O (δ

2H: -26.2 ± 4.4 ‰, δ

18O: -2.8 ±

0.9 ‰), compared with water in perennial reaches (δ2H: -53.8 ± 0.8‰, δ

18O: -7.7 ±

0.1‰)(ANOVA: δ2H: F(1,53) = 92.6, p < 0.001, δ

18O: F(1,53) = 74.65, p < 0.001;

Figure 2.2), confirming the greater influence of evaporation at intermittent sites and

likely disconnection from alluvial groundwater. In contrast, most designated

perennial sites lay lower on the LEL, suggesting groundwater inflow. There was no

overall discernible longitudinal gradient in isotopic composition (i.e. progressive

enrichment) from the Upper to Lower catchment. However, if intermittent and

perennial reaches were considered separately, δ2H and δ

18O did reveal a pattern of

evaporative loss across downstream gradients (intermittent: δ2H: R

2 = 0.21, p =

0.016, δ18

O: R2 = 0.152, p = 0.045; perennial: δ

2H: R

2 = 0.62, p < 0.001, δ

18O: R

2 =

0.44, p < 0.001; Figure 2.2b, 2.2c), even though there was considerable variability

across intermittent reaches.

Dissolved inorganic carbon concentration ranged from 3.53 to 84.86 mg L-1

(median:

71.63 mg L-1

) during April 2014. DIC concentrations (mean ± standard error) were

significantly higher at perennial sites (77.4 ± 0.9 mg L-1

) than intermittent sites (51.8

± 4.6 mg L-1

) (ANOVA: F(1,17) = 22.19, p < 0.001). However, there was no

significant difference in DIC concentration between Lower and Upper catchments.

The stable isotope ratio of dissolved inorganic carbon (δ13

CDIC) did not differ

between perennial and intermittent sites (F(1,53) = 0.12, p = 0.728) and ranged from -

13.02 to -5.04 ‰ (median: -10.70 ‰). There was no significant difference in δ13

CDIC

between Lower and Upper catchments. Although across the whole catchment,

δ13

CDIC in perennial reaches became progressively more depleted downstream, yet

this pattern was not present in intermittent reaches (Figure 2.3).

28

Figure 2.2 a) Stable isotope composition of perennial (red) and intermittent (yellow) stream

waters, located in Lower (□) and Upper (○) catchment positions in the Fortescue River

catchment, northwest Australia. Also shown are rainwater (+) and groundwater () samples

collected from the region at the time of this study. The local evaporation line (LEL) was

calculated from 55 surface water samples whilst the local meteoric water line (LMWL) was

sourced from Dogramaci et al. (2012). b) Relationship between the distance from river

mouth and stable hydrogen (δ2H) and c) oxygen (δ

18O) isotopes of water.

29

Total dissolved nitrogen concentrations were highly variable, ranging from 40 to

1391 µg L-1

across all sites (median: 220 µg L-1

), and did not differ between

perennial and intermittent reaches overall (F(1,53) = 0.5, p = 0.48). TDN

concentrations were significantly higher in the Lower- than the Upper- catchment

(ANOVA: F(1,53) = 27.20, p < 0.001), and there was a significant longitudinal

increase in TDN (R2 = 0.169, p = 0.002, Figure 2.3). Soluble reactive phosphorus

concentration in the water column was also highly variable and ranged from < 1 to

39.4 µg L-1

(median: 2.7 µg L-1

). SRP was significantly higher in intermittent

reaches (6.0 ± 1.4 µg L-1

) compared to perennial reaches (2.7 ± 0.3 µg L-1

)

(ANOVA: F(1,53) = 9.98, p = 0.002), although there was no overall longitudinal trend

in SRP concentrations across the whole catchment (Figure 2.3). The N:P ratio of the

water column thus ranged from 3 to 563 (median: 78)(Redfield, 1934), suggesting

potentially broad scale P limitation to pool productivity across both the Lower and

Upper catchments. There was a significant longitudinal increase in N:P ratio (R2 =

0.182, p < 0.001, Figure 2.3). N:P ratios were significantly lower in intermittent

compared to perennial reaches (ANOVA: F(1,53) = 6.768, p = 0.012). Four reaches

had N:P ratios < 16:1, all of which were known highly ephemeral pools of

Coondiner Creek in the Upper Fortescue Catchment (Fellman et al., 2011).

Dissolved organic carbon concentrations ranged from 0.78 to 34.01 mg L-1

across the

whole catchment (median: 2.64 mg L-1

). DOC was significantly higher in

intermittent reaches (4.9 ± 1.2 mg L-1

) than perennial reaches (3.2 ± 0.7 mg L-1

)

(ANOVA: F(1,53) = 5.069, p = 0.029). DOC concentrations increased downstream

across these perennial reaches (R2 = 0.162, p = 0.034), while this pattern was not

present across intermittent reaches (Figure 2.3). SUVA254 ranged from 0.06 to 3.16 L

30

mg-1

m-1

(median: 1.19 L mg-1

m-1

), corresponding to 0 to 25 % aromaticity (median:

11.4 %). Overall there was no significant difference in SUVA254 values between

perennial and intermittent reaches, between Lower and Upper catchments, nor a

discernible longitudinal gradient.

31

Figure 2.3 Relationship between the distance from river mouth and total dissolved nitrogen

(TDN), soluble reactive phosphorus (SRP), N:P ratio, dissolved organic carbon (DOC),

dissolved inorganic carbon (DIC), and the stable isotope composition of dissolved inorganic

carbon (δ13

CDIC). Shaded fill denote perennial (red) and intermittent (yellow) stream waters,

symbols denotes Lower (□) and Upper (○) catchment position (n = 55).

32

PARAFAC decomposition of EEM spectra identified three fluorescing DOM

components in the dataset (Figure 2.4). Component C1 had excitation maxima of

250, 345, 445 nm with emission maxima 464 nm. Component C2 had excitation

maxima of 250 and 310 nm and emission maxima of 394 nm. Components C1 and

C2 were matched in the OpenFluor library with terrestrially derived humic-like

components C1 and C2 from Shutova et al. (2014) (r2

= 0.99 and 0.98). C1

terrestrially delivered organic matter, and C2 reprocessed terrestrially delivered

organic matter. Component C3 had excitation maxima of 250 and 280 nm and

emission maxima 304 nm. Component C3 was matched with the tyrosine-like C5

from Yamashita et al. (2013) (r2 = 0.92). Examination of error residuals from the

three component PARAFAC model indicated that a tryptophan-like component was

also present in some samples from the catchment, although the core consistency of a

four component model was lower and the model unable to be split-half validated

with our dataset. Overall there was no significant difference in DOM Fmax values for

components C1, C2, or C3 between persistent and intermittent streams, between

Lower and Upper catchments, nor discernible longitudinal gradients.

33

Principal component analysis indicates that hydrochemistry of streams is most likely

related to stream hydrology. The first three axes of the PCA explain 65 % of the total

variance of hydrochemical variables across the catchment (Figure 2.5). Water stable

isotopes δ2H and δ

18O contribute consistently to the spread of sites along PC1, which

Figure 2.4 a) Excitation-emission spectra of the three modelled components from

PARAFAC analysis. b) Spectral loadings for excitation (red dash) and emission (blue line)

wavelengths for each component. Components 1 and 2 are humic-like, Component 3 is

protein-like tyrosine. c) Relationship between the stable isotope composition of water (δ18

O)

and dissolved organic matter components. Shaded fill denote perennial (red) and intermittent

(yellow) stream waters, symbols denotes Lower (□) and Upper (○) catchment position (n =

55).

34

explained 27.1 % of the variance. Perennial and intermittent streams generally

grouped apart across this axis (Figure 2.5). The humic-like dissolved organic matter

components C1 and C2 contribute to the spread of sites along PC2, which explained

23.4 % of the variance.

2.3.2. How is stream nutrient availability linked to hydrology across the

catchment?

In order to explore how nutrient availability and organic matter in surface water are

related to hydrologic characteristics, dissolved nutrients (N and P) and dissolved

organic and inorganic carbon were plotted against the δ18

O signatures (Figure 2.6).

Figure 2.5 Principal component analysis of hydrochemical parameters of perennial (red) and

intermittent (yellow) stream waters, symbols denotes Lower (□) and Upper (○) catchment

position (n = 55). The first two axis of the PCA explains 50.5 % of the total variance. Water

stable isotopes δ2H, δ

18O and δ

13C-DIC contribute consistently to PC1, dissolved organic

matter components C1, C2, and contribute to PC2.

35

SRP was positively correlated, albeit weakly, with δ18

O across the catchment (R2 =

0.16, p = 0.003, n = 55). DOC was also weakly correlated with δ18

O across the

catchment (R2 = 0.16, p = 0.003, n = 55) for all streams combined. The strength of

the correlation is somewhat leveraged by one point for each of these parameters,

although there is no a priori reason to exclude the site as an outlier. δ13

CDIC was also

positively correlated with δ18

O across the catchment (R2 = 0.31, p < 0.001, n = 55).

In contrast, DIC concentration was negatively correlated with δ18

O (R2 = 0.44, p =

0.002, n = 19). TDN and N:P ratios of the water column were also related to degree

of pool evaporation in intermittent reaches, where both TDN (R2 = 0.23, p = 0.010, n

= 26) and N:P ratio (R2 = 0.13, p = 0.006, n = 26) tended to increase at more

enriched δ18

O signatures. In contrast, DOM in perennial but not intermittent reaches

was positively correlated to δ18

O values: DOM Fmax values for components C1 (R2 =

0.17, p = 0.032, n = 28), C2 (R2 = 0.17, p = 0.031, n = 28), and C3 (R

2 = 0.23, p =

0.010, n = 28) (Figure 2.6).

36

Figure 2.6 Relationship between the stable isotope composition of water (δ18

O) and total

dissolved nitrogen (TDN), soluble reactive phosphorus (SRP), N:P ratio, dissolved organic

carbon (DOC), dissolved inorganic carbon (DIC), and the stable isotope composition of

dissolved inorganic carbon (δ13

CDIC). Shaded fill denote perennial (red) and intermittent

(yellow) stream waters, symbols denotes Lower (□) and Upper (○) catchment position (n =

55).

37

2.4. Discussion

This study demonstrates the diversity of hydrological drivers and biogeochemical

characteristics of surface waters that can occur across dynamic catchments in hot

arid environments. This diversity stems primarily from stream reach isolation and

variable connectivity to alluvial water, and reflects the continuum of hyporheic flow

paths that can occur across the catchment. Perennial and intermittent reaches showed

distinct differences in hydrology that in turn resulted in high variability in nutrient

concentration and stoichiometryacross sites. This study also revealed stream

hydrology to be a stronger determinant of stream chemistry across the catchment,

although individual chemical parameters differed in their behaviour between

intermittent and perennial sites. Below, I discuss patterns in stream chemistry along

the upstream-downstream longitudinal gradient, followed by patterns across an

evaporative gradient throughout the Fortescue River catchment.

2.4.1. Longitudinal gradients

The chemical composition of surface water throughout the Fortescue River

catchment showed no clear patterns. At the catchment scale, one of the few

longitudinal gradients present was for nitrogen with both TDN concentration and

N:P ratio increasing from upstream to downstream. No other parameters measured in

this study indicated a gradient from upstream to downstream. (Vannote et al.,

1980)The downstream increase in TDN across the Fortescue catchment may reflect

an increase in the relative contribution of groundwater (a NO3- source) to streams

downstream. Hence, rather than instream processes (i.e. nutrient spiralling and

denitrification), nitrogen in intermittent systems is more governed by subsurface and

lateral biogeochemical processes.

38

When looking at surface waters of perennial and intermittent reaches discretely it

became clear that the hydrochemistry of these two reach types were governed by

separate processes. Across both perennial and intermittent reaches the degree of

evaporative loss progressively increased from upstream to downstream, although

evaporative loss was much greater in intermittent reaches than perennial reaches

across the catchment. Again, this is showing that flow cessation and pool isolation

create key differences in drivers of biogeochemistry between intermittent and

perennial reaches. As pools lose connectivity to alluvial groundwater, more localised

factors such as waterbody geometry (esp. surface area to volume ratio) along with

degree of exposure/shading to wind and solar irradiation will strongly determine

evaporative loss. Hence as pools become progressively isolated across catchments of

intermittent systems the remnant longitudinal stream processes get overridden by

stronger hydrological connectivity and evaporation processes.

An upstream to downstream carbon gradient was present only across perennial

reaches. DOC concentrations, along with DOM humic- and protein-like components

increased from upstream to downstream across these reaches. Whilst DIC

concentrations did not show any gradient, its isotopic ratio (δ13

CDIC) became more

negative downstream. In Mediterranean IRES DOM becomes increasingly

heterogeneous throughout the catchment during dry/drought periods, whereas DOM

is relatively homogeneous during flood (Ejarque et al., 2017). Overall, I found no

downstream carbon gradients for the Fortescue River. While not directly addressing

DOC concentrations, the RCC predicts that the diversity of dissolved organic

compounds decreases downstream, and CPOM:FPOM ratio decreases (Vannote et

39

al., 1980). There have been some recent developments in conceptualising DOM

dynamics in river catchments, although they have done little to address flow

intermittency. The Pulse-Shunt (Raymond et al., 2016), and Active-Pipe (Casas-Ruiz

et al., 2017) concepts have been developed to characterise perennial systems. While

downstream patterns of carbon were indeed evident across perennial reaches in this

study, this does not hold true for intermittent reaches of the Fortescue River

catchment. Hence the inclusion of flow intermittency and discontinuum should be a

focus of future revisions of these models.

2.4.2. Evaporative loss drives stream hydrochemistry

This study demonstrates that broad scale processes across the Fortescue River

catchment may be more appropriately framed by hydrological parameters such as the

degree of evaporation (i.e. a hydrological gradient, rather than a geographic

longitudinal gradient). Flow intermittence and hydrological connectivity in IRES

naturally creates a river dis-continuum once flow cessation has occurred and reaches

become isolated. Hence, longitudinal gradients in intermittent catchments may at

best show a legacy ‘ghost’ of prior flows, and do not well describe the ‘now’

conditions of the catchment, especially in dryland regions such as the Pilbara where

‘dry’ inter-flow periods are long compared to other IRES globally. For example the

evaporation : inflow ratio derived from water stable isotopes explains patterns in

catchment biogeochemistry across an intermittent tropical catchment (Fellman et al.,

2014). As hydrology is an important determinant of stream dynamics overall, a

gradient based on hydrological character (such as δ2H and δ

18O values) may be more

useful for understanding biogeochemical processes in IRES.

40

The evaporative pressure on stream reaches (based on δ18

O values) was related to

both carbon and phosphorus content across the catchment. DOC and SRP

concentration increase as pools contract due to evaporation. In dryland streams of

regions where flow is short-lived and rates of evaporation are high – compared to

other IRES where isolation is primarily driven by flow gradually ceasing – pools

become isolated and contract and evapo-concentration of nutrients and carbon

becomes a significant process (Hamilton et al., 2005; Sheldon & Fellows, 2010;

Siebers et al., 2016). This concentration of materials may maintain production as

these pools contract. The evaporative pressure on stream reaches also showed

patterns in stream nitrogen, although these were only present across intermittent

reaches. TDN concentrations and N:P ratios increase as pools contract due to

evaporation. This is also seen in the increase in TDN downstream across the overall

catchment, whereby stream water is progressively evaporated as it transits

downstream, and evapo-concentration acts as a nutrient retention mechanism

(McLaughlin, 2008). Consequently, we may expect rates of in-stream productivity to

be higher in lowland streams than their upland counterparts throughout the

catchment.

The isotopic δ13

CDIC values measured within the catchment indicated that the bulk of

dissolved inorganic carbon in these streams is sourced from regional groundwater

aquifers. Once DIC enters surface waters it has an inverse relationship with the

evaporative pressure on stream reaches. DIC concentration decreases as pools

contract due to evaporation, whilst δ13

CDIC values increase. The relationship between

δ13

CDIC and DIC concentration in stream water is indicative of DIC source (Miller &

Tans, 2003; Campeau et al., 2017). In this case the linear slope (δ13

CDIC: -0.07 ‰)

41

indicates a geogenic source such as a carbonate bearing aquifer. Groundwater may

hence be a major source of inorganic carbon to stream food webs in the Pilbara

region via these DIC inputs (Fellman et al., 2014). The pattern seen here seems

counterintuitive when comparing to evapo-concentration of other dissolved solutes

(i.e. DOC, TDN and SRP), although it demonstrates that either DIC is degassing

(Cao et al., 2016), being utilised for photosynthesis (Madsen & Sand-Jensen, 1991;

Chen et al., 2016), or both. For example, charophytes in carbonate rich waters are

known to take-up and precipitate bicarbonate excess to their metabolic needs (Kufel

et al., 2013). In both cases (degassing or metabolic uptake) the lighter 12

C isotope in

DIC is preferentially fractionation resulting in an increase in δ13

CDIC value.

Consequently, these results make sense in that DIC is ‘used up’ as overall

evaporative pressure increases.

This study indicates that isolation and evaporative pressure are stronger factors than

longitudinal continuum (i.e. RCC) in shaping stream chemistry across the Fortescue

River catchment, and presumably intermittent systems in general. Other processes

such as sediment sorption, nutrient uptake and organic matter cycling not accounted

for would also play a role in local variations. Further investigation into the

biogeochemical processes and uptake of nutrients and carbon in these streams is

required to understand why we observed only a weak correlation between pool

hydrology and hydrochemical parameters. Whilst it is important to understand the

spread (boundary conditions) and variance in dissolved organic matter specific for

the region, there is perhaps a scaling issue between what we see at site and at

catchment scale (e.g. von Schiller et al., 2017). To improve our understanding of

these scaling issues we would require a more comprehensive spatial and temporal

42

sampling effort across the catchment. Whilst surface waters which persist well

beyond flow cessation in intermittent systems may be hot spots in stream

biogeochemistry (McClain et al., 2003), we have sampled ‘cool moments’. This is a

period when much more local processes take over from catchment wide events.

Pilbara streams are unique and diverse habitats in an arid landscape. The Fortescue

River catchment houses streams with diversity in biogeochemical nature which is

primarily driven by evaporation and disconnection from alluvial water.

43

3. PHOSPHORUS SORPTION AND DISSOLVED ORGANIC

MATTER INTERACTIONS IN IRON-RICH STREAM

SEDIMENTS

3.1. Introduction

Phosphorus (P) sorption kinetics of stream sediments plays a key role in regulating P

content of the water column and thus has a strong influence on ecological processes

within aquatic systems (Reddy et al., 1999). Competition for limited P resources

between biotic and abiotic components of freshwater streams is increased when

sediments are dominated by minerals with high sorption capacity (Fink et al., 2016).

However, organic matter (OM) content may also influence P sorption kinetics of

stream sediments (McDowell & Sharpley, 2001), albeit in inconsistent and complex

ways. For example, P sorption in aquatic systems is linked to iron cycling and

governed by the so-called “Mortimer’s ferrous wheel” (Mortimer, 1941). What this

means is that under anaerobic conditions OM enhances the microbially-mediated

reductive dissolution of iron (III) minerals and subsequently releases the phosphate

ion (Mitchell & Baldwin, 1998; Watts, 2000a). In contrast, aerobic oxidation of iron

(II) back to iron (III) produces amorphous iron (III) oxyhydroxides with high binding

capacity, which in turn enables the recovery (sorption) of a portion of the P released.

Aerobic sorption of P to ferric mineral surfaces may be diminished by dissolved

organic matter (DOM) competing for sorption sites (Wang et al., 2008; Verbeeck et

al., 2017). The mechanisms that underpin differences in P-OM interactions in

freshwater ecosystems are complex yet likely play a key role in determining overall

nutrient limitation of within-stream productivity.

44

Allochthonous inputs from riparian vegetation can make a significant contribution to

sediment OM loads, as well as contribute to dissolved OM (DOM) in the water

column, particularly in intermittent streams of dryland regions (Baldwin, 1999;

Fellman et al., 2013; Datry et al., 2018b). Litter (leaves and small branch material),

as well as larger coarse woody debris, may enter reaches and isolated pools directly

from overhanging riparian vegetation or can be washed into streams from the

surrounding catchment during rainfall and flood events (Gonçalves et al., 2014;

Tonin et al., 2017). Direct inputs of litterfall from overhanging riparian vegetation is

especially important in dryland regions where the stream corridor – and particularly

around more persistent pools – is much more productive than the surrounding

catchment. This tannin rich litter releases significant amounts of DOM in to the

water column through the leaching of fresh litter, as well from decomposing

material. In intermittent and ephemeral streams this DOM can then be further

concentrated as surface water contracts during drought (Siebers et al., 2016; Harjung

et al., 2018).

DOM is known to influence P sorption to sediments in a number of ways. For

example, DOM can inhibit phosphate sorption to sediments through the occupation

and blocking of mineral surface sites (Gu et al., 1995). Fe catalysed photo-

degradation of DOM produces more bioavailable fractions of DOM as well as Fe(II),

which can be rapidly oxidised to Fe(III), producing amorphous Fe(III)

oxyhydroxides (Howitt et al., 2008). DOM might also act as a chelating agent in

iron-rich sediments capable of interacting with both the electron acceptors on

sediment surfaces ‘freeing’ iron and phosphate into solution (Baken et al., 2016;

45

Daugherty et al., 2017). These possible interactions are further complicated by the

variable nature of the DOM, which in freshwater ecosystems is comprised of several

different fluorescing components, including humic and fulvic acids, tryptophan and

tyrosine. Recent studies have also shown that the concentration and composition of

DOM of ephemeral streams – including the Pilbara – is strongly influenced by

surface/alluvial-water hydrodynamics (Fellman et al., 2011; Siebers et al., 2016;

Harjung et al., 2018). Thus we might expect that both hydrology and the

composition (fluorescence characteristics) of DOM will impact P sorption/desorption

processes in sediments of ephemeral and intermittent streams.

Ephemeral and intermittent streams are further characterised by cycles of drying and

rewetting, which are known to release more P than sediments or soils that are

consistently moist (Grierson et al., 1998) but also decrease the P affinity of

streambank sediments compared to sediments that remain permanently inundated

(Watts, 2000b). Drying of sediments can also decrease P adsorption capacity

(Attygalla et al., 2016), while P desorption and release from sediments is enhanced

upon rewetting (Baldwin, 1996; Kerr et al., 2010). These effects can persist for

extended periods of time. For example, P adsorption capacity was decreased for

several weeks after rewetting in wetland clays and silts (Song et al., 2007). Baldwin

and Mitchell (2000) also showed that wetland sediments that go through decadal

wetting-drying cycles have reduced adsorption capacity even when the current

inundation period has spanned many prior years. There is also high heterogeneity in

exposure to the extent of drying and rewetting that occurs within ephemeral and

intermittent streams owing to varying levels of connectivity to alluvial groundwater

(Fellman et al., 2011; Siebers et al., 2016). We might thus expect that stream

46

sediments in reaches that are less connected to groundwater and therefore dry out,

have reduced P adsorption characteristics than sediments that remain persistently

inundated. These combined processes of drying/rewetting cycles in association with

increased DOM concentrations in individual pools and different forms of DOM,

allow for the production of P-OM-Fe complexes, which may enable P to remain

suspended in solution as colloidal particles (Baken et al., 2016; Yan et al., 2017), or

form more complex structures to sorb to sediment mineral surfaces (Cheng et al.,

2004). Consequently, shifts in flood and drought cycle and heterogeneity in the

distribution of surface and alluvial groundwater along streams will have significant

impacts on biogeochemical cycles and in-stream processes (Leigh et al., 2015).

This chapter investigated phosphorus sorption processes and how DOM may modify

P sorption, in iron-rich sediments from pools along an intermittent Pilbara stream. In

this study, P sorption/desorption kinetics of sediments were compared with and

without additions of DOM derived from allochthonous sources, and among pools of

contrasting hydrological connectivity to the alluvial water table. Pools were defined

as either ‘persistent’, where alluvial through flow maintains water level in the pool

throughout the year, or ‘ephemeral’, where pools are disconnected from the alluvial

groundwater and hence sediments frequently dry out in the weeks to months

following cease to flow. I hypothesised that: i) sediments from ‘ephemeral’ pools

will have lower P adsorption capacity than sediments from ‘persistent’ pools; ii) an

increase in DOM concentration will decrease the P adsorption capacity of sediments;

and iii) an increase in DOM concentration will increase P desorption from

sediments.

47

3.2. Methods

3.2.1. Study site and sampling

Sediments and leaf litter samples were collected from six pools along 8 km of

Coondiner Creek (-23.00° S, 119.62° E), an intermittent (ephemeral) stream in the

Hamersley Ranges of northwest Australia (Figure 3.1). Coondiner Creek is

considered representative of many streams in the region that cut through a network

of semi-confined gorges throughout the ranges (Fellman et al., 2011; Siebers et al.,

2016). Stream bed sediments are primarily reworked channel iron deposits,

containing a high proportion (50 – 60%) of Fe-hydroxide minerals, primarily

hematite and goethite sourced from banded iron formations in the catchment

(Ramanaidou et al., 2003). Throughout the catchment there are also minor granitic

bodies and carbonate deposits (dolerite and calcrete) around springs (Skrzypek et al.,

2013).

48

Pools were classed as persistent (n = 3) or ephemeral (n = 3) as defined above based

on previous hydrological studies of the stream (Fellman et al., 2011), and their

hydrologic status at the time of sampling confirmed via analysis of stable isotope

ratios of surface waters. Sediments (0-5 cm depth) were collected from pools using a

hand corer in June 2016. Sediments were collected from three positions around each

pool at ~ 40 - 50 cm water depth and generally within 2 m of the pool edge. Cores

were bulked by pool and then passed through a 2 mm sieve, with the < 2 mm

fraction retained and sealed in plastic ‘zip-lock’ bags. Wet sediments were packed in

a portable cooler for transport to the laboratory. All sediments were then oven-dried

(30 °C) for 72 h in preparation for adsorption-desorption experiments.

Figure 3.1 a) The Fortescue river catchment (solid fill) of the semi-arid Pilbara region

(hatching), northwest Australia. b) Location of Coondiner Creek in the Upper Fortescue

River catchment, c) ‘persistent’ (black squares) and ‘ephemeral’ (grey circles) pools sampled

along Coondiner Creek.

49

Leaf litter of the most dominant riparian tree species, Eucalyptus camaldulensis

subsp. refulgens, was collected adjacent to the pools to prepare DOM extracts. Litter

leachates were prepared by gently agitating 20 g leaf material in 1 L MilliQ water for

2 h on a shaker table (Fellman et al., 2013). The leachate was then filtered (Whatman

GF/F 0.7µm) to exclude particulate organic matter and retain the dissolved

components. The DOM leachate was prepared the day before batch experiments

were performed. Leachates were stored in the dark wrapped in foil and refrigerated

prior to use.

3.2.2. Sediment mineralogy and elemental chemistry

Sediment pH and electrical conductivity (EC) was measured in a 1:10 (w/v) soil-

solution. The pH was measured in the laboratory with a bench top pH meter (Orion

model 520A). A subsample of each sample was dried (50 °C) and ground to a fine

powder (< 60 µm) for mineralogical and chemical analysis. OH-extractable

inorganic (OH-Pi) and OH-extractable total P (OH-Pt) were measured after

extraction in 0.1 M NaOH (Bowman & Cole, 1978). Phosphorus concentrations in

all extracts were measured using the modified ascorbic acid method (Kuo 1996;

Murphy and Riley, 1962). Sediment crystalline structure and amorphous content was

determined by quantitative X-ray diffraction (XRD) (PANalytical Cubix3, Almelo,

Netherlands). Samples were packed and presented as un-orientated powder mounts.

ZnO and CaF2 internal standards were added to each sample and amorphous content

was determined by an internal standard scan. Sediment chemical characterisation

was determined by X-ray fluorescence (XRF) by Intertek Genalysis (Perth, WA)

using standard techniques. A fused disk of the powdered sample was prepared with

50

borate flux. Percent loss on ignition (%LOI) was determined after combustion at 425

°C, 650 °C, and 1000 °C by thermogravimetric analysis.

3.2.3. Phosphorus sorption characteristics

Phosphorus (P) adsorption experiments were undertaken by adding a range of known

P concentrations to sediments (Pant & Reddy, 2001; Nair & Reddy, 2013). In the

samples with no additions of DOM (DOM0), 5 g sediment was weighed into 125 mL

high-density polyethylene (HDPE) bottles containing 100 mL of phosphate solution

(0, 2.5, 5, 10, 20, 40, or 100 mg L-1

in 0.003 M KCl). Solution blanks were included

in all steps of the procedure to assess P recovery and ensure P was not sorbed to the

HDPE bottle. Samples were mixed on an end-over-end shaker at 22 °C (± 2 °C) for

24 h. Sediments were allowed to settle for 30 min prior to the supernatant being

filtered (Sartorius minisart 0.45µm). P concentrations in the filtered supernatant were

then analyzed using a modified ascorbic acid method (Murphy & Riley, 1962; Kuo,

1996).

To investigate the effects of dissolved organic matter (DOM) on P adsorption

characteristics, the above procedure was repeated on two further subsets of

sediments by adding leachates with (i) DOC concentrations of 3.7 mg L-1

(DOM5),

or (ii) DOC concentrations of 44.6 mg L-1

(DOM50). The two DOC treatment

concentrations were selected based on observed water column DOC concentration in

the system (Siebers et al., 2016), and a 10-fold increase to emulate the expected

higher DOC concentrations encountered in sediment pore water. P concentrations of

the filtered supernatant were analysed as above. The differences among treatments in

51

the optical absorption and DOM fluorescence properties was investigated on a subset

of samples (Ci = 0, 40, 100 mg L-1

).

The amount of P sorbed was calculated as:

𝑞𝑒 =𝑉(𝐶𝑖−𝐶𝑒)

𝑊 [Equation 1]

Where qe is the ratio of P sorbed by sediment (mg g-1

), Ci and Ce are the initial and

equilibrium P concentrations (mg L-1

), V is the solution volume (L) and W is the

weight of sediment (g). Plotting the amount of P sorbed (qe) vs equilibrium

concentration (Ce) we examined the linear region of the curve (Ce < 10 mg L-1

):

𝑞𝑒 = 𝐾𝑙𝐶𝑒+𝑏 [Equation 2]

Here, the slope is equal to the linear adsorption coefficient (Kl), and intercept (b)

equal to initial soil P. The equilibrium P concentration (EPC0) was then calculated:

𝐸𝑃𝐶0 =𝑏

𝐾𝑙 [Equation 3]

The experimental data was fitted to both Langmuir and Freundlich isotherm models

using non-linear least squares. The Freundlich equation is as follows (Freundlich,

1907):

𝑞𝑒 = 𝐾𝐹𝐶𝑒1/𝑛

[Equation 4]

Where KF (L mg-1

) is a constant related to adsorption energy, and n is a correction

factor.

The Langmuir equation (Langmuir, 1918) used was:

𝑞𝑒 =𝐾𝐿𝑏𝐶𝑒

1+𝑏𝐶𝑒 [Equation 5]

Where KL (L mg-1

) is the Langmuir isotherm constant, and b (mg g-1

) is the

maximum adsorption capacity, often referred to as ‘Smax’ (Pant & Reddy, 2001).

52

P desorption was measured at the conclusion of the incubations. For each of the three

treatments (DOM0, DOM5, and DOM50) the supernatant was discarded and the

sediment retained. Then, 100 mL 0.003 M KCl was added to each bottle and the

sediment shaken for 24 h. Samples where then filtered (0.45 µm) and the supernatant

analysed for P in solution as previously described.

Dissolved organic matter (DOM) characterisation

DOM remaining in solution for a subset of P treatments following batch adsorption

experiments was characterised in order to assess if particular components of DOM

were more likely to be sorbed than others. These subset treatments with initial P

concentration (Ci) of 0, 40 and 100 mg P L-1

were further analysed by absorbance

and fluorescence spectroscopy. Absorbance spectra 200 to 800 nm were measured on

a UV-Visible spectrophotometer (Varian Cary 50 Probe). Dissolved organic carbon

(DOC) and total dissolved nitrogen (TDN) were measured simultaneously on a

Shimadzu TOC-V analyser coupled with a total nitrogen module (Shimadzu TNM-

1). C:N ratios of the equilibrated solutions were calculated as DOC:TDN. Specific

UV absorbance at 254 nm (SUVA254) (Weishaar et al., 2003) and spectral slope

(S275-295) (Helms et al., 2008 ) were calculated from the absorbance spectra.(Helms et

al., 2008 )

DOM fluorescence was measured on a Varian Cary Eclipse spectrofluorometer

(Varian Medical Systems, Inc. California USA). Extracts were diluted to within

optical range with an auto-dilutor (Hamilton, Microlab) to avoid self-quenching due

to inner-filter effects (Ohno, 2002). Excitation emission spectra (EEM) were

53

produced for excitation wavelengths 240 to 450 nm at 5 nm intervals with emission

intensities captured from 300 to 600 nm at 2 nm intervals. Humification index (HIX)

was calculated following the methods outlined in Ohno (2002). The main

contributing fluorescing organic matter components of resulting EEM’s were

extracted and quantified via parallel factor analysis (Stedmon & Bro, 2008; Murphy

et al., 2013).

3.2.4. Data analyses

Non-metric multidimensional scaling (nMDS) of sediment characteristics to assess

similarity among pools was performed following normalisation, and quantified with

analysis of similarities (ANOSIM) in Primer (Primer v 6.1.18). Sorption model

fitting and statistical analysis were performed in R version 3.4.1 (R Core Team,

2017). Adjusted R-squared values and residual sum of squares (RSS) were assessed

to support sorption model selection. Parallel factor analysis (PARAFAC) was

performed in MATLAB (R2012a) using the n-way and drEEM (v4.0) toolboxes

(Murphy et al., 2013). The PARAFAC model was trained to look for best fit between

three to seven fluorophore components within the experimental data. The four-

component model was validated using split-half analysis (Stedmon & Bro, 2008),

model components were compared to other fluorophore components in the

OpenFluor spectral library (Murphy et al., 2014) with OpenChrom (v1.3.0 Dalton).

Peaks are presented as maximum fluorescence intensity (Fmax) values in Raman units

for each component. Finally, differences in DOM components and indices with pool

hydrology, DOM treatment, and initial P concentration were investigated with

PERMANOVA routines following log(x + 1) normalisation of Fmax values

(PERMANOVA+ v 1.0.8).

54

3.3. Results

3.3.1. Sediment properties

A summary of sediment characteristics is provided in Table 3.1. Overall, sediments

collected from pools along Coondiner Creek were iron-rich and similar in character

irrespective of hydrological regime (i.e., there was no difference between ‘persistent’

versus ‘ephemeral’ pools). The OH-extractable inorganic P (Pi) (0.2 – 1.0 µg P g-1

)

and total P (PT) (70 – 82 µg P g-1

) were also similar among pools (Table 3.1), with

the exception of ‘Pool w’. This pool had elevated P content (Pi: 4.38, PT: 302 µg P g-

1) and in contrast to the other pools sampled, was located adjacent to the main

channel, and therefore received runoff from a small tributary catchment. Elemental

Figure 3.2 Non-metric multidimensional scaling (nMDS) plots of a) elemental composition

of sediments from XRF, and b) mineralogy of sediments from XRD for ‘persistent’ (black

squares) and ‘ephemeral’ (grey circles) pools of Coondiner Creek. Data were normalised

prior to scaling.

55

XRF showed sediments were predominantly composed of iron and aluminium

oxides (Fe: 26 – 42 %, Al2O3: 5 – 10 %). XRD analysis showed sediment

mineralogy was primarily composed of hematite (20 – 37 wt %) and goethite (10 -

16 wt %), along with quartz (11 – 23 wt %). XRD results were plotted in nMDS

space to further assess if there were differences in sediment chemistry among pools

or with different hydrology (Figure 3.2). Analysis of similarity between sediments

from ‘persistent’ and ‘ephemeral’ pools indicated that sediment geochemistry did not

differ with pool hydrology; dissimilarities between groups were no greater than

dissimilarities within groups for both XRF chemical characterisation (ANOSIM:

global-R = -0.074, P = 0.70), and XRD crystalline and amorphous content

(ANOSIM: global-R = -0.037, P = 0.50).

56

Tab

le 3

.1 C

hem

ical

ch

arac

teri

stic

s o

f se

dim

ents

coll

ecte

d f

rom

Coondin

er C

reek

. E

lect

rica

l co

nd

uct

ivit

y (

EC

) an

d p

H w

ere

mea

sure

d i

n a

1:1

0 (

w/v

) so

il-s

olu

tio

n.

Bulk

sed

imen

t sa

mple

s fr

om

eac

h p

ool

wer

e ai

r-dri

ed p

rior

to c

hem

ical

an

alysi

s.

57

Figure 3.3 Fluorescent DOM components derived from fluorescence spectroscopy and

PARAFAC analysis. a) Modelled excitation-emission spectra of humic-like components 1

and 2, protein-like component 3, and unknown component 4. b) Excitation (red dash) and

emission (blue line) spectral loading of each corresponding component.

58

Ta

ble

3.2

Fre

un

dli

ch a

nd L

angm

uir

model

par

amet

ers

fitt

ed t

o e

xper

imen

tal

adso

rpti

on

iso

ther

ms.

KF:

Fre

un

dli

ch

adso

rpti

on

en

ergy c

on

stan

t, n

: F

reundli

ch c

orr

ecti

on f

acto

r, K

L:

Lan

gm

uir

iso

ther

m c

on

stan

t (L

mg

-1),

b:

Lan

gm

uir

max

imu

m a

dso

rpti

on

cap

acit

y (

mg g

-1).

Mea

n v

alues

wit

h s

tandar

d d

evia

tion i

n p

aren

thes

is (

n =

3).

Mo

del

fit

s w

ere

com

par

ed u

sin

g a

dju

sted

-R2,

and r

esid

ual

sum

of

squar

es (

RS

S).

59

3.3.2. DOM properties of litter leachates

PARAFAC decomposition of EEM spectra identified four fluorescing components in

the DOM dataset (Figure 3.3). Component C1 had excitation maxima of 310 and 400

nm with emission maxima 420 – 444 nm. Component C2 had excitation maxima of

255, 340, and 445 nm and emission maxima of 464 – 502 nm. Components C1 and

C2 were matched in the OpenFluor library with humic-like components C1 and C2

from Baker et al. (2014) (r2

= 0.93 and 0.92). Component C3 had excitation maxima

of 280 and 390 nm and emission maxima 306 – 356 nm. Component C3 was

matched with the tryptophan-like C5 from Baker et al. (2014) (r2 = 0.90).

Component C4 had excitation maxima of 250 nm and emission maxima of 304 and

was matched with tyrosine-like C6 from Shutova et al. (2014) (r2 = 0.93). The leaf

litter leachates used in the experiments consisted of humic-like C1 and C2 peaks and

a tryptophan-like C3 peak. Tyrosine-like C4 was absent from the leachate.

3.3.3. Phosphorus sorption characteristics

Phosphorus sorption characteristics for sediments from all pools were generally best

described by the Freundlich model (Table 3.2; Figure 3.4), although there were some

exceptions where a Langmuir model provided a better fit (e.g., persistent pools were

more similar to the Langmuir model when DOM was present; Table 3.2). Results for

both Freundlich and Langmuir isotherms are thus reported for comparison.

The Freundlich adsorption capacity KF ranged from 0.024 to 0.049 mg g-1

. There

was no difference in KF values between DOM treatments or with pool hydrology.

‘Persistent’ and ‘ephemeral’ pools also did not differ in sorption characteristics

(Table 3.2). However, the Freundlich n constant was significantly higher in the

60

DOM5 treatment (2.364 – 3.157) compared to both the DOM0 (1.705 – 2.192) and

DOM50 (1.952 – 2.278) treatments (Two-way ANOVA: P = 0.002).

The Langmuir model generally fitted adsorption data where Ce < 50 mg L-1

, but was

less accurate at higher concentrations and thus underestimated maximum sorption

capacity (Figure 3.4). The Langmuir isotherm constant (KL) ranged from 0.386 to

1.012 L mg-1

(Table 3.2). There was no difference in KL values between DOM

treatments or hydrology. Langmuir adsorption capacity (b) ranged from 0.106 to

0.152 mg g-1

. P adsorption capacity was significantly lower in the DOM5 treatment

(0.106 – 0.138 mg g-1

) compared to DOM0 (0.141 – 0.149 mg g-1

) and DOM50

(0.114 – 0.152 mg g-1

) (Two-way ANOVA: P = 0.004). ‘Persistent’ and ‘ephemeral’

pools did not differ in sorption characteristics and there was no interaction between

hydrology and DOM.

Figure 3.4 Experimental data from batch phosphorus adsorption experiments fitted to

Freundlich (solid line) and Langmuir (dash) isotherms. Mean adsorption (qe) and standard

error (n = 3) for sediments from ‘persistent’ (black square) and ‘ephemeral’ (grey circle)

pools versus equilibrium P concentration (Ce) shown.

61

3.3.4. Desorption of P from iron-rich sediments

The amount of P desorbed after 24 h ranged from 0 to 0.05 mg g-1

across the three

DOM treatments, demonstrating that much of the added P is retained by the Fe-rich

sediment once adsorbed (Figure 3.5). However, P desorption from sediments also

increased with the amount of P previously adsorbed. Increasing the amount of DOM

in the sediment-water mix generally increased the amount of P desorbed from

sediment. A linear fit to the experimental data indicated that the amount desorbed

was 10.0 – 12.7 % (Pearson’s r = 0.98, 0.87) for the DOM0 treatment, 11.7 – 16.4 %

(r = 0.35, 0.49) for DOM5, and 17.2 – 23.9 % (r = 0.69, 0.95) for DOM50. However,

when DOM5 was added, nearly all of the P sorbed up to 0.1 mg g-1

was retained by

the sediment (Figure 3.5). In this case, significant desorption occurs only once the

sediment P concentration exceeded 0.1 mg g-1

.

Figure 3.5 Phosphorus adsorption (Pads) versus desorption (Pdes) patterns of Pilbara

sediments. Values shown are means with standard error (n = 3) of sediments from

‘persistent’ (black square) and ‘ephemeral’ (grey circle) pools.

62

3.3.5. Changes in DOM composition with incubation and P adsorption

As expected, DOM in solution at the beginning and end of 24 h incubation in the

DOM0 treatment was below detection limit. Hence, carbon based parameters are

reported only for DOM5 and DOM50 treatments (Figure 3.6). Interestingly, the DOC

concentration remaining in solution following batch adsorption experiments was

higher for ‘persistent’ pools than ‘ephemeral’ pools in both DOM5 and DOM50

treatments. However, there was no clear relationship between initial P concentration

(Ci) and final DOC concentration. There was a positive relationship between the C:N

ratio of the equilibrated solution and initial P concentration up to 20 mg P L-1

for the

DOM50 treatment. At initial P concentrations above 20 mg L-1

, the C:N ratio of the

equilibrated solution did not change (Figure 3.6). SUVA254 values following batch

adsorption experiments ranged from 0.35 to 1.20 L mg-1

m-1

. The SUVA254 value

increased with increasing initial P concentration for the range 0 to 20 mg P L-1

for

both the DOM5 and DOM50 treatments (Figure 3.6).

Absorption spectra indicate that the composition of DOM remaining in solution at

the end of the incubation correlates to the initial P concentration (Figure 3.7). The

spectral slope value (S275-295) increased with increasing P concentration (Figure 3.7).

There was an initial rapid increase in slope value as initial P increased from 0 to 10

mg L-1

with spectral slopes flattening out at higher initial P concentrations. The

humification index (HIX) ranged from 0.46 to 0.81 (Figure 3.7).

63

Figure 3.6 Dissolved organic carbon (DOC), C:N ratio, and Specific UV absorbance at

254 nm (SUVA254), at the conclusion of batch phosphorus adsorption experiments.

Values shown are means with standard error (n = 3) for sediments from ‘persistent’

(black square) and ‘ephemeral’ (grey circle) pools. Note different scales on y-axis

between DOC panels.

64

PARAFAC analysis of the fluorescence EEM spectra split the experimental dataset

into four DOM components. The fluorescence maxima for all DOM fluorescing

components were reflective of the DOC concentration of the initial DOM treatment

additions (Figure 3.8). Fmax values were significantly different between DOM

treatments (F2,45 = 56.82, P = 0.001), with DOM50 > DOM5 > DOM0. Component

correlation plots whilst validating the model, also indicated that Fmax values for

humic-like C1 and C2 components were highly correlated (Pearson’s r = 0.94).

Fluorescence maxima (Fmax) at the end of 24 h batch phosphorus adsorption

experiments for sediments from ‘persistent’ and ‘ephemeral’ pools indicated that

humic acids (components C1 & C2) were desorbed from sediments, particularly

when initial P concentration was high (Figure 3.8). Fluorescence maxima for these

components significantly increased with increasing initial P concentration (C1: F2,45

= 4.88, P = 0.006, C2: F2,45 = 23.61, P = 0.001). There was also an interaction

between initial P and DOM treatment (C1: F4,45 = 4.55, P = 0.003, C2: F4,45 = 7.99,

P = 0.001). Tryptophan-like C3 fluorescence maxima did not change with increasing

initial P concentration (F2,45 = 0.09, P > 0.05) and there was no interaction between

initial P and DOM treatment for this component (F4,45 = 2.18, P > 0.05). Tyrosine-

like C4 appears to be either a degradation product or extracted OM from the

sediments as this component was not present in the initial DOM leachates (Figure

3.8). Fluorescence maxima significantly increased with increasing initial P

concentration (F2,45 = 4.71, P = 0.017) and there was a significant interaction

between initial P and DOM treatment (F4,45 = 11.4, P = 0.001).

65

Figure 3.7 UV-vis and fluorescence indices measured at the conclusion of batch phosphorus

adsorption experiments. Spectral slope (S275-295) and humification index (HIX) values are

presented as means with standard error (n = 3) for sediments from ‘persistent’ (black square)

and ‘ephemeral’ (grey circle) pools.

66

Figure 3.8 Fluorescence maxima (Fmax) for PARAFAC derived DOM components at the

end of 24 h batch phosphorus adsorption experiments for sediments from ‘persistent’ (black)

and ‘ephemeral’ (grey) pools. X-axis indicates the initial P concentration (Ci) in batch

experiments and y-axis indicates fluorescence maxima (Fmax) of DOM components.

Components 1 and 2 are humic-like fluorophores, component 3 is protein-like (amino acids),

whilst component 4 is thought to be a sediment derived OM degradation product. Values are

given as mean and standard error (n = 3).

67

3.4. Discussion

The results of this study demonstrate that phosphorus adsorption and desorption

processes are both important for mediating the concentrations of potentially

biologically ‘available’ phosphorous in the water column in dryland streams. The

results also illustrate the complex ways that DOM may interact with iron-rich

sediments to influence P availability in the water column. While the iron-rich

sediments examined in this study act as a sink for phosphorous in the water column,

the concentration of DOM in the water column also appears to play a role in

regulating P desorption. In particular, DOM increased P desorption from sediments

when sediment P loading was high (i.e. Pads is above saturation). However when

sediment P loadings were high, DOM desorption from sediments also increased.

Consequently, it appears that P displaces DOM from surface sites on these

sediments.

Overall the adsorption isotherms from the sediments from Coondiner Creek are

classical "Type 1" monolayer forms, which have an initial steep slope up until

saturation is exceeded and then reaches equilibrium as the sediment adsorption

capacity is approached (Brunauer et al., 1938). Since Type 1 curves are typically

monolayer; P and DOM appear to be primarily competing for limited surface

adsorption sites in the sediment. While I measured phosphorus adsorption capacity

on relatively coarse-textured (1 – 2 mm) sediments, adsorption was comparable to

lateritic soils of much finer texture (Singh and Gilkes (1991); Zhang et al. (2012).

However, maximum adsorption capacity measured in these Pilbara sediments of 100

– 150 mg kg-1

was not high as might be expected for sediments so high in Fe (Kerr et

68

al., 2011), but did fall within range of other stream sediment studies (Agudelo et al.,

2011).

3.4.1. DOM composition is influenced by the presence of excess P

This study demonstrates that along with phosphorus, these stream sediments are a

sink for ‘recalcitrant’ DOM. The preferential removal from solution of DOM

exhibiting high spectral slope (S275-295) when P was absent or at low experimental

concentrations (< 10 mg L-1

), coincides with P saturation for the stream sediments

used in this study. While this study did not directly measure DOM molecular weight,

correlation between spectral slope and molecular weight has been observed

previously (Helms et al., 2008; Wagner et al., 2015; Wunsch et al., 2018).

Furthermore, high molecular weight DOM is preferentially sorbed to goethite (Ohno

et al., 2007), a key mineral in the sediments of Coondiner Creek. The analyses of

decreased SUVA254 (Figure 3.6), spectral slope (S275-295), humification index (HIX)

(Figure 3.7), and fluorescing humic-like components (Figure 3.8) suggests that DOM

retained in solution at lower P concentrations was of decreased aromaticity. Thus, it

is likely that aromatic high molecular weight (HMW) DOM is being preferentially

adsorbed on to sediments. If reduced aromaticity was a result of P bonding to DOM

in solution and breaking conjugated bonds then the opposite trend (i.e. decreased

aromaticity at higher initial P concentrations) would be expected. These results are

consistent with other studies that have investigated DOM adsorption in isolation

from the effects of P, showing increased molecular weight of DOM remaining in

solution with increasing DOM concentration (Gu et al., 1995; Meier et al., 1999).

69

Phosphorus addition enhanced desorption of native DOM from the Fe-rich sediments

examined. Fluorescence analysis revealed that humic acids (components C1 & C2)

were desorbed from sediments (Figure 3.8), most notably when initial P

concentration was high. However this was not discernible from DOC analysis alone,

as DOC concentrations were below detection limit. This finding suggests that P

replaces DOM associated with sediment surfaces, causing the release of stored DOM

back to the water column. Increasing phosphorous concentrations has been shown to

enhance the of release DOM from peat sediments (Sokolowska et al., 2011). The

presence of organic matter has also been shown to limit desorption of inorganic P

from eutrophic lake sediments through preferential desorption of DOP (Wang et al.,

2008). Consequently, in pools of intermittent and ephemeral streams, where litter

inputs can be significant (Datry et al., 2018b), an increase in P to the system from the

cumulative impact of mining and agriculture in the region may interfere with carbon

cycling by altering the rate of remineralisation of sediment OM. This would be most

strongly felt in pools at the more ephemeral end of the hydrological spectrum.

3.4.2. Surface/alluvial hydrodynamics do not control sediment P sorption at the

pool scale

A surprising result of this experiment was that P sorption did not differ significantly

between contrasting pool hydrologies. However, XRD and XRF analysis

demonstrated that sediments were of a similar bulk chemical and mineralogical

nature regardless of pool hydrology (Figure 3.2). This consistency in sediment type

across the 8 km of creekline examined here is likely due to regular reworking during

high discharge events that homogenises the sediments at scales greater than

individual pools. However, the frequent exposure of sediments in the field to both

70

high temperatures (where soils and dry stream beds are extremely hot for extended

periods – often in excess of 60 °C throughout the day and for weeks on end) as well

as subject to prolonged drying and rewetting cycles may have increased the

crystallinity of iron minerals (Attygalla et al., 2016) to a point where iron

crystallinity across both ‘persistent’ and ‘ephemeral’ pools becomes similar.

Attygalla et al. (2016) showed that substantial changes to the iron mineralogy in fine

and organic rich sediments was evident at drying even at 30 °C. Interestingly the

max P adsorption for sediments observed in the coarse mineral Pilbara sediments

here was very similar (~ 0.15 mg g-1

) to sediments dried at temperatures of 30 °C (or

higher) in Attygalla et al. Further experimentation to compare undried versus dried

sediments from a range of sites and mineralogies will help elucidate whether a lack

of difference among sites observed here is truly reflective of site conditions or

possibly also in part an artefact of sample treatment.

It would be useful to gain a better understanding of what the timeframe of ‘resetting’

of adsorption capacities is. As far as I am aware, no other study has investigated

post-rewetting ‘reset time’ or the effect of historical frequency of wet/dry cycles on

sediment sorption kinetics, although the issue was indirectly explored by Baldwin et

al. (2000). Over periods of years to decades, the sorption characteristics of sediments

of streams that have ephemeral wet/dry hydrological regimes may act differently to

streams with perennial flow. Consequently, as stream intermittency increases

globally (Acuña et al., 2015), sediment processes in these streams will be affected by

longer duration of flooding or indeed more regular dry periods. A consequent change

in sediment P sorption properties may be one mechanism that in turn impacts on

stream nutrient cycling and metabolism (e.g. Acuña et al., 2015; Sabater et al.,

71

2016). Further investigation of the relative importance of nitrogen and phosphorus

limitation of producers in drylands streams will enable us to better understand the

ecological outcomes of changing sediment P sorption properties.

Overall, the findings of this study have helped elucidate the fate of P in a highly

dynamic, and notionally oligotrophic, dryland catchment. I hypothesised that the

differences in wet/dry cycling and hydrological residence time in contrasting

‘persistent’ and ‘ephemeral’ pools would affect P sorption behaviour. Other studies

investigating P dynamics have previously shown drying affects both adsorption and

desorption characteristics of sediments (Baldwin, 1996; Watts, 2000b). My original

hypothesis was based on a model where sediment is spatially static and the

adsorption capacity of sediments increases with duration of inundation (i.e. time of

sampling was n months after last flood). However, my results revealed was no

evidence of differences between ‘persistent’ and ‘ephemeral’ pools in sediment P

sorption, even though ephemeral pools are more prone to drying out annually, and

sediments are exposed to ambient air temperatures exceeding 45 °C. Drying itself

can affect the crystalline structures of Fe-rich minerals (Attygalla et al., 2016). One

explanation is that high energy flood events transport and mix dry and wet sediments

and thus ultimately determine the over-riding sorption of P from the catchment onto

colloidal Fe and other fines in suspension (Müller et al., 2006; Baken et al., 2016).

Therefore it is likely that floods are an important ‘P adsorption period’, whilst quiet

periods post-cessation of flow are important for the concentration of DOM and

desorption of P to the water column. These shifts in the relative importance of

different components of biogeochemical dynamics at different times in the

72

hydrologic cycle would help explain the second flush of productivity often seen in

ephemeral pools as they dry-down (Siebers et al., 2016).

73

4. DOES LOW PHOSPHORUS LIMIT THE SHORT-TERM

METABOLIC RESPONSE OF PHYTOPLANKTON AND

CHAROPHYTES OF INSTREAM POOLS ON AN

INTERMITTENT DRYLAND STREAM?

4.1. Introduction

Like many freshwater aquatic ecosystems worldwide, autotrophic production of

intermittent rivers and ephemeral streams (IRES) in hot arid regions is limited by

nutrient availability, particularly of nitrogen (N) and phosphorus (P) (Grimm &

Fisher, 1986; Tank & Dodds, 2003; von Schiller et al., 2011). Phytoplankton,

charophytes and other macrophytes are ubiquitous to autochthonous primary

production within streams. In IRES, the productivity of phytoplankton has been

found to be linked to discharge patterns and seasonal flows, owing to associated

changes in water quality, including turbidity, dissolved carbon and nutrient

availability (Townsend & Douglas, 2017). However, in smaller streams with more

gravelly sediments, the water column may be highly transparent, such that

charophytes and submerged macrophytes may be more significant to overall

production, similar to what has recently been observed for oligotrophic shallow lakes

(Martinsen et al., 2017). Consequently, nutrient limitation of productivity of IRES

are likely to vary considerably from pool to pool in periods of low or no flow

depending on relative dominance of phytoplankton versus charophytes and

associated pool conditions.

74

Dryland streams around the world have been primarily considered nitrogen limited

(e.g. Grimm & Fisher, 1986), in part because most early studies were focussed on

relatively young geologic landscapes in the northern hemisphere (Grimm et al.,

1981; Tank & Dodds, 2003). In contrast, the intermittent streams of the semi-arid

Pilbara region of northwest Australia are considered to be oligo- to meso-trophic but

are especially depauperate in phosphorus (P) (see Chapter 2). Low P availability in

these streams has been predominantly attributed to the ancient geology of the Pilbara

craton (ca. 2.4 bya; Kranendonk et al., 2002; Arndt et al., 2007), such that the

lithosphere has been depleted in P via weathering processes and/or occluded into

secondary minerals (Wild, 1958). P availability in the water column is further

diminished due to geochemical reactions via adsorption to Fe-rich sediments

(Chapter 3). Consequently, P limitation of autotrophic production within these

streams is likely.

Intermittent and ephemeral streams (IRES) in arid regions pose a novel environment

in which to test questions of nutrient limitation on primary production. In addition to

being highly P impoverished – and also frequently low in nitrogen (McIntyre et al.,

2009b; Pinder et al., 2010) – Pilbara streams are characterised as having highly

variable summer flows and are therefore extremely intermittent in nature (Kennard et

al., 2010). Storms and large rainfall events, which may exceed 100 mm in a day,

drive catchment recharge (Dogramaci et al. 2015), although these storms and their

impacts can be highly variable both spatially and temporally (Ruprecht, 1996;

Rouillard et al., 2015). In the prolonged dry periods between flows, many of these

streams retract to isolated pools, some of which are maintained by alluvial

groundwater (persistent pools; Chapter 3). These pools may remain in this somewhat

75

isolated state for months to years (Fellman et al., 2011). Hence, in the otherwise

xeric landscape, these isolated pools are biogeochemical and ecological hotspots

between flow events, acting as refugia for fish (Morgan & Gill, 2004) and supporting

aquatic plant and invertebrate communities as well as terrestrial fauna.

Charophytes are green algae that have structure superficially similar to higher plants

and are distributed across many aquatic habitats in northern Australia (Masini, 1988;

Casanova, 2005b; Pinder et al., 2010), where they form dense beds that act as

physical structure for epiphytic communities and are important for fish grazing and

habitat (Schneider et al., 2015). Charophyte beds also act as nutrient sinks in aquatic

systems (Kufel & Kufel, 2002; Rodrigo et al., 2007). Charophytes are well adapted

to growth in oligotrophic systems (Kufel & Kufel, 2002), partly because they are

able to take up available nutrients from the sediment as well as directly from the

water column (Littlefield & Forsberg, 1965; Andrews, 1987; Komuro et al., 2017;

Rodrigo et al., 2017). This physiological trait allows charophytes to out-compete

phytoplankton for available P and maintain ‘clear’ water (characteristic of

charophyte dominated systems) by sequestering nutrients in their tissues. Hence,

high abundance of charophytes in a system would perpetuate phytoplankton biomass

remaining low. However, in ephemeral pools, charophyte mats become a source of

nutrients as pools contract and dry; nutrients are subsequently released into streams

on rewetting (Lu et al., 2017). Consequently, as pools re-fill after stream flow events,

competition occurs between charophytes and phytoplankton to establish whether a

clear charophyte-dominated or turbid phytoplankton-dominated system develops.

76

To date, the overall extent of nutrient limitation of phytoplankton and charophytes

within ephemeral Pilbara streams have not been directly tested. However, there is

now an urgent need to better understand the potential responsiveness of these

systems to nutrient additions in order to assess the escalating and likely cumulative

impacts of development across the region from intensification of pastoral production

through large-scale irrigation projects (www.agric.wa.gov.au/r4r/irrigation-pilbara),

preferential grazing of livestock in riparian corridors (Lyons, 2015) as well as

discharge of groundwater that is potentially enriched in nitrate from mine dewatering

(e.g. Halse et al., 2014). Based on pool stoichiometry, previous studies have alluded

to Pilbara streams being primarily P-limited (Pinder et al., 2010; Fellman et al.,

2011) or N and P co-limited (McIntyre et al., 2009a; McIntyre et al., 2009b).

However, estimations of response of primary production to nutrient additions

provide more direct evidence of the effects of nutrients individually or in

combination.

As previously discussed (Chapter 2, Chapter 3), connectivity to alluvial water can

also have a profound effect on organic matter inputs and nutrient fluxes (Fellman et

al., 2011; Siebers et al., 2016). Therefore in this chapter the metabolic responses of

primary producers to additions of nitrogen and/or phosphorus were compared among

pools of contrasting hydrology. Specifically, I used diel dissolved oxygen curves to

examine rates of ecosystem metabolism between ‘persistent’ and ‘ephemeral’ pools

along an arid zone stream. Concurrently I used bottle assays to assess the rate of

carbon uptake of phytoplankton and charophytes, and to test for nutrient limitation of

phytoplankton and charophytes with additions of N and/or P.

(Puche et al., 2018)

77

4.2. Methods

4.2.1. Site description

The focus for this study was Coondiner Creek, an intermittent dryland stream

situated in the Pilbara region of Western Australia (see Chapter 3, Figure 3.1). Field

experiments were conducted at Coondiner Creek during a dry-phase in August 2015.

Flow events in Coondiner Creek are rare and short lived and the dry-phase – which

may last seasons or multiple years – is the common state for the creek. At the time of

sampling, surface flows had ceased several months prior and the creek had retracted

to a series of isolated pools along the main channel. Pools were classed by their

hydrology as either predominately ‘persistent’ or ‘ephemeral’ based on previous

studies of pool-alluvium connectivity within the study catchment (Fellman et al.,

2011; Siebers et al., 2016). Connectivity to groundwater was confirmed for this

study from analysis of water stable isotope ratios measured at the time of sampling.

4.2.2. Pool water physicochemistry of persistent and ephemeral pools

Water samples were collected for analysis of stable isotopes (δ2H, δ

18O, δ

13CDIC) and

nutrients (NOx, NH4, SRP). Samples for δ13

CDIC, δ2H and δ

18O analyses were filtered

through a sterile syringe filter (PALL 0.2 µm Supor) into a glass vial ensuring all

headspace was removed. Conductivity and pH were measured in the field using a

handheld probe (YSI model 85). Water isotope and DOC/TDN samples were

immediately refrigerated (4 °C) in the field and stored until analysis. Water samples

for nutrient analysis were filtered through a sterile syringe filter (Sartorius mini-sart

0.45 µm) and frozen in the field until time of analysis. Water samples were analysed

for pool water physicochemistry as outlined in previous chapters.

78

4.2.3. Estimation of net ecosystem production

Net ecosystem production (NEP) of each pool was calculated using the single station

dissolved oxygen method (Odum, 1956; Staehr et al., 2010). Data loggers were

deployed at each pool for between 24 to 72 h to measure dissolved oxygen (DO) and

temperature (YSI ProODO). Sensors were placed at the approximate centre of pool

and ~30 cm below the surface. DO and temperature were measured and logged at a

15 minute intervals. Metabolic parameters (GPP, NEPdaytime, CR, and NEM) were

calculated following Staehr et al. (2010). Measurement periods were very still and

wind had a minimal effect on the pools over the duration of the observations; hence

piston velocity (k600 = 0.0214 m h-1

) was estimated based on a nominal wind speed

of 0.5 m s-1

.

4.2.4. Nutrient limitation experiments

Bottle incubation experiments were undertaken in-situ to assess nutrient limitation

within the water column (phytoplankton) and on charophytes (Chara sp.). Plastic

300 mL biological oxygen demand (BOD) bottles (Environmental Express,

Charleston, South Carolina) were used for the incubation experiments. Five

treatments (Light, Dark, +N, +P, +N+P), with three replicate bottles of each

treatment per site were used to assess responses to nutrient additions (procedures

described further, below). Water for the experiment was collected from each pool

and amended with 13

C-enriched sodium bicarbonate (13

C, 99%) (Cambridge Isotope

Laboratories Inc. Andover, MA) to make a final solution of 0.536 mg C L-1

above

ambient concentrations and an isotopic enrichment of +2000 ‰ over ambient DIC

values (mean ambient = ~60 mg C L-1

DIC and -12.6 ‰ δ13

CDIC; see Table 4.1 for

individual site values). A water sample was taken immediately before and once the

79

water had been enriched to confirm actual δ13

CDIC values for each series of

incubations.

Tab

le 4

.1 A

mb

ient

dis

solv

ed n

itra

te/n

itri

te (

NO

x),

am

mo

niu

m (

NH

4),

an

d s

olu

ble

rea

ctiv

e phosp

horu

s (S

RP

) co

nce

ntr

atio

ns

of

stre

am w

ater

at

each

pool.

N:P

rat

ios

calc

ula

ted a

s th

e ra

tio b

etw

een D

IN a

nd

SR

P w

her

e D

IN =

NO

x +

NH

4.

Val

ues

giv

en

are

mea

ns

and

sta

ndar

d d

evia

tion (

n =

3).

Am

bie

nt

dis

solv

ed i

norg

anic

car

bo

n (

DIC

) co

nce

ntr

atio

n a

nd

its

car

bo

n i

soto

pe

rati

o (δ

13C

DIC

) ar

e al

so g

iven

.

80

Charophytes were hand collected from each pool and gently agitated to remove loose

detritus but retain attached epiphyton. Sections of charophyte thalli were cut below

the fourth node and two of these were added to each bottle. A short piece of silicone

tubing was used to deliver the 13

C-enriched site water to the bottom of each BOD

bottle to ensure aeration of the water was minimised while filling each bottle. Bottles

were amended with either nitrate (prepared from KNO3) and/or orthophosphate

(prepared from KH2PO4) to make a final concentration of 500 µg N l-1

and/or 50 µg

P l-1

above ambient pool water concentrations (Table 4.1) to obtain a ten-fold

increase in both N and P concentrations in the water column. Bottles were

subsequently capped, the time recorded and placed in pool to incubate under ambient

conditions for 4 - 6 h. At the conclusion of the incubation, bottles were collected

from the pool and immediately processed for isotopic sampling. Each 280 mL of

incubated bottle water was filtered through a pre-combusted and pre-weighted quartz

fibre filter (Whatman QMA-25, particle retention 2.2 µm) and phytoplankton

retained on the filter. Filtrate was retained for analysis to determine NO3, NH4 and

PO4 concentrations at the end of incubation. MilliQ water (50 mL) was then passed

through the filter to flush any remaining 13

C-enriched water out of the filter matrix.

Filters were wrapped in aluminium foil and placed in individual ziploc bags.

Charophyte thalli were removed from the bottle, rinsed with MilliQ water, blotted

dry with lint-free tissue (Kimwipes) and placed in individual ziploc bags. All

samples were immediately frozen in the field until analysis in the laboratory.

4.2.5. Laboratory analyses of N, P and carbon

Charophytes and quartz filters containing phytoplankton were oven dried at 60 °C

for 24 h. Charophyte samples were weighed to estimate dry biomass then ground to a

81

fine powder in a ball mill (Retsch MM200) in preparation for bulk stable isotope

analysis. Ground charophyte samples were split and one half acidified with 4 % HCl

for 4 h to remove carbonates. Samples were then rinsed three times in MilliQ water,

oven dried and ground. The remaining non-acidified half was retained for separate

stable isotope analysis of δ15

N as the acidification process alters δ15

N values (Bunn

et al., 1995; Mazumder et al., 2010). Quartz filters were also acidified to remove any

carbonates/inorganic carbon.

Dissolved organic carbon (DOC) and total dissolved nitrogen (TDN) of water

samples were measured simultaneously on a Shimadzu TOC-V analyser coupled

with a total nitrogen module (Shimadzu TNM-1). Nitrate (NO3) and Ammonia

(NH4) were measured by continuous flow analysis (Technicon Autoanalyzser II).

Dissolved inorganic nitrogen (DIN) was calculated as DIN = NO3 + NH4. Soluble

reactive phosphorus (SRP) was measured spectrophotometry by the molybdenum

blue method (Murphy & Riley, 1962). pH was measured with a benchtop pH meter

(Orion model 520A).

4.2.6. Stable isotope analysis of plant tissues, filters and water samples

Carbon and nitrogen isotopes of solid samples (plant tissues and quartz samples)

were measured on an isotope ratio mass spectrometer with elemental analyser

(Thermo Delta V). Stable isotope values were reported in delta (δ) notation relative

to the Vienna Pee Dee Belemnite for δ13

C, and atmospheric nitrogen for δ15

N

(equation 1).

𝑅𝑝𝑒𝑟𝑚𝑖𝑙 = (𝑅𝑠𝑎𝑚𝑝𝑙𝑒

𝑅𝑠𝑡𝑎𝑛𝑑𝑎𝑟𝑑− 1) × 1000 [Equation 1]

82

Carbon isotope of dissolved inorganic carbon (δ13

CDIC) and DIC concentration were

measured on a Thermo Delta XL IRMS with Gasbench II. Water isotopes (δ2H and

δ18

O) were measured on a Picarro Cavity Ring-Down Spectrometer. Multi-points

normalization was used in order to reduce raw values to the international scale

(Skrzypek, 2013). Technical details of the instrument and used procedure can be

found in the introduction of Skrzypek and Ford (2014). All isotope values are given

in per mil [‰, VSMOW] according to delta notation (Coplen, 1996).

4.2.7. Calculation of productivity based on uptake of 13

CDIC

Rate of uptake of 13

C by charophytes (Chara sp) and phytoplankton were estimated

from the change in their isotopic compositions over the period of the incubation and

then converted to rates of production (P, mg C g-1

DW h-1

) following Nayar et al.

(2009). Briefly;

𝑃(𝐶ℎ𝑎𝑟𝑎) = 𝐶∗(𝐶𝑒𝑡−𝐶𝑏𝑡)

𝑊∗𝑡∗(𝐶𝑒𝑤−𝐶𝑏𝑡) [Equation 2]

where: P = productivity (mg C g-1

DW h-1

)

C = total carbon content of tissue (mg)

W = dry weight of tissue (g DW)

t = duration of incubation (h)

Cet = 13

C in the enriched tissue (atom %)

Cbt = 13

C in the background tissue (atom %)

Cew = 13

C in the enriched water (atom %)

𝑃(𝑃ℎ𝑦𝑡𝑜𝑝𝑙𝑎𝑛𝑘𝑡𝑜𝑛) = 𝑃𝑂𝐶∗𝑉∗(𝐶𝑒𝑡−𝐶𝑏𝑡)

𝑊∗𝑡∗(𝐶𝑒𝑤−𝐶𝑏𝑡) [Equation 3]

83

where: POC = total carbon content of tissue (mg)

V = volume of BOD bottle (L).


Statistical procedures were conducted in R (R Core Team, 2017) and PRIMER 6

(Primer-E Ltd. UK). Principal component analysis (PCA) was first performed on

pool environmental variables to confirm their hydrological status and assess for any

variance in environmental conditions during the incubations. Environmental data

were normalised then square root transformed prior to analysis. Isotopic and

productivity data was checked for normality (Shapiro-Wilk test) and homogeneity of

variance (Bartlett’s test) prior to executing a two-way ANOVA with nutrient

treatment and pool hydrology as factors. Statistical tests were assessed to be

significant at α = 0.05. Post-hoc Tukey’s HSD was performed for comparisons

where statistically significant differences within factors occurred.

84

Ta

ble

4.2

C:N

:P s

toic

hio

met

ry o

f pool

wat

er,

char

ophyte

s an

d p

hyto

pla

nkto

n.

85

4.3. Results

4.3.1. Pool hydrology and water chemistry

Principal component analysis confirmed that pools operationally defined as

‘persistent’ or "ephemeral’ were clearly differentiated based on their hydrochemistry

(Figure 4.1). The first two principal components explained 68.6 % of the variance

(39.8 % by PC1 and 28.8 % by PC2) (Figure 4.1). SIMPER analysis indicated that

difference between persistent and ephemeral pools was best explained by water

Figure 4.1 PCA ordination diagram of ‘persistent’ (black squares) and ‘ephemeral’ (grey

circles) pools of Coondiner Creek and environmental variables. Cond: Electrical

conductivity , TSS: total suspended solids, TDN: total dissolved nitrogen, δ2H: water stable

isotope deuterium, DIC: dissolved inorganic carbon, 13C-DIC: δ13

CDIC, DOC: dissolved

organic carbon, DO(avg): average dissolved oxygen, DO(range): dissolved oxygen range, NH4:

ammonium, Temp(avg): average water temperature, Temp(range): Water temperature range,

NOx: nitrate/nitrite, SRP: soluble reactive phosphorus, Chl a: Chlorophyll a

86

isotopic ratios (δ2H, δ

18O), DIC concentration and pH. DIN:SRP ratios were < 10:1

for all Coondiner Creek pools at the time of the experiment (Table 4.2). The N:P

ratios of charophyte tissues collected from pools were all less than 6:1 (Table 4.2).

4.3.2. Ecosystem metabolism

Measured dissolved oxygen in the pools ranged from 1.84 to 16.98 mg L-1

, with diel

peak to peak amplitude within any one pool from 2.19 to 10.85 mg L-1

(Figure 4.2).

Surface water temperature ranged from 12.9 to 24.3 °C across all pools, with diel

peak to peak amplitude from 1.2 to 6.3 °C. Gross primary production (GPP) in

Coondiner creek pools estimated from diel dissolved oxygen curves ranged from

1.40 to 20.64 g O2 m-3

d-1

. Community respiration (CR) ranged from 1.16 to 18.64 g

O2 m-3

d-1

. GPP:CR ratios for all pools were > 1, indicating that pools were net

autotrophic at the time of the study (Figure 4.2).

87

Fig

ure

4.2

Die

l dis

solv

ed o

xygen

curv

es f

or

stre

am p

ools

alo

ng C

oondin

er C

reek

. V

alu

es f

or

gro

ss p

rim

ary p

rod

uct

ivit

y (

g O

2 m

-3

d-1

) (G

PP

), c

om

mu

nit

y r

espir

atio

n (

CR

24),

net

eco

syst

em p

roduct

ion (

NE

P),

and G

PP

:CR

rat

io a

re g

iven

on

the

figu

re f

or

each

po

ol.

Gre

y a

nd

lig

ht

shad

ing i

ndic

ate

nig

ht

and d

ay p

erio

ds.

Boxed

are

as s

ignif

y t

ime

envel

op

e fo

r b

ott

le a

ssay

s.

88

4.3.3. 13C enrichment due to photosynthetic uptake of

13C-enriched HCO3

The δ13

C values of phytoplankton were compared at the end of the incubation period

(4 – 6 hours) in both dark and light treatments for bottles both with and without 13

C-

HCO3 additions in order to estimate enrichment due to photosynthetic uptake (Figure

4.3). Phytoplankton incubated without 13

C-HCO3 additions had δ13

C values of -35.12

+/- 1.19 ‰ when incubated under dark conditions and -33.07 +/- 1.10 ‰ under light

conditions. Phytoplankton incubated in presence of 13

C-HCO3 additions had similar

δ13

C values of -34.53 +/- 2.22 ‰ when incubated under dark conditions (controls)

but were highly enriched 103.75 +/- 19.64 ‰ under light conditions.

89

4.3.4. Short-term metabolic response of phytoplankton and charophytes to nutrient

enrichment

Chara tissues and phytoplankton were significantly enriched in 13

C after incubation

with labelled 13

C-HCO3. δ13

C values were significantly higher for ‘light’ treatments

compared to the ‘dark’ treatments for both charophytes (F = 63.7, p < 0.001) and

phytoplankton (F = 51.65, p < 0.001), confirming uptake through photosynthesis

(Figure 4.3). Based on changes in carbon isotope composition through the

incubation, photosynthetic production (P) by charophytes in the control (unamended)

treatments was estimated as 2.14 +/- 0.8 mg C g-1

DW hr-1

. Two-way ANOVA of P

showed that neither nutrient treatment (F = 0.72, p = 0.546) or pool hydrology (F =

0.03, p = 0.874) significantly influenced charophyte production. Hence, charophyte

primary production was not limited by either dissolved N or P. There was also no

interaction effect between nutrient treatment and pool hydrology (F = 0.47, p =

Figure 4.3 Carbon stable isotope ratios of phytoplankton under light/dark conditions

incubated in situ with and without 13C-enriched HCO3 added. Bars are means and error

bars indicate standard error (n = 3).

90

0.703). Charophyte tissue nitrogen content and productivity measured at the end of

incubation were positively and significantly correlated (R2 = 0.65, p < 0.001; Figure

4.5). Productivity of charophytes were also positively correlated to phosphorus tissue

content (R2 = 0.40, p < 0.001) and tissue N:P ratio (R

2 = 0.22, p < 0.001) (Figure

4.5).

Figure 4.4 Short-term productivity response of charophytes and phytoplankton to nutrient

additions in ‘persistent’ and ‘ephemeral’ pools estimated as rate of 13C-enriched HCO3

uptake. Bars are means and error bars indicate standard error (n = 3).

91

Photosynthetic production (P) by phytoplankton in the control (unamended)

treatments was 0.017 +/- 0.01 mg C g-1

DW hr-1

. Photosynthetic production (P) by

phytoplankton was also higher in ‘persistent’ pools compared to ephemeral pools (F

= 9.84, p = 0.003; Figure 4.2). There was no significant increase in productivity with

nutrient addition (F = 1.24, p = 0.302) and no interaction effect between nutrient

treatment and pool hydrology (F = 0.07, p = 0.976).

4.4. Discussion

This study indicated that both charophytes and phytoplankton were metabolically

unresponsive to increases in nutrient availability in the short-term (< 6 h). The

results indicated that both nitrogen and phosphorus were not limiting in the short-

term. Nutrient additions did not significantly increase, nor supress, the rate of

production over the experiment period. It was somewhat surprising that

phytoplankton did not show a short-term response to spikes in N and P, as the

simpler single cell thickness structure of planktonic algae would presumably enable

faster response to environmental perturbations. Other studies have shown an

increase of photosynthetic rate of charophytes in presence of elevated nitrogen

Figure 4.5 Relationship between the rate of production and tissue content in charophytes at

the end of the incubation experiment. Charophyte content of a) nitrogen, (%) b) phosphorus,

and c) nitrogen:phosphorus (N:P) ratio.

92

(Hough & Putt, 1988) or phosphorus (Francoeur, 2017), but these assays were

conducted over 12 - 48 hours, allowing more time for charophytes to respond to

nutrient additions than this study. Conversely, over longer timeframes (i.e. several

weeks), nitrate (Lambert & Davy, 2011) and phosphate (Forsberg, 1964) have also

been shown to have an inhibitory rather that stimulatory effect on growth of

charophytes. Phosphorus concentrations greater than 20 µg PO4-P L-1

can also

impact negatively on charophyte growth (Wetzel, 2001). In contrast, phytoplankton

blooms are a common occurrence when P concentrations are elevated as a result of

eutrophication. However, it is likely that N versus P limitation will vary with

seasonal factors that reflect changes in hydrology, connectivity and particularly

temperature. Recent studies have also shown that charophyte responses to nitrogen

additions are temperature dependent (Puche et al., 2018). For example, warming

favours increased %N and thus N:P ratio of the Mediterranean charophytes Chara

vulgaris and C. hispida (Puche et al., 2018). Pool temperatures varied among pools

in this study. While the effect of temperature on nutrient uptake was not formally

tested it may better explained differences between pools than differentiation of pools

by hydrologic status.

Resolving the charophyte taxonomy beyond genus level within Coondiner Creek

would be useful in determining if any of the variation in response observed between

pools is due to species differences. The interactions between nitrogen uptake and

temperature may also be dependent on species; pioneer species may exhibit more

phenotypic plasticity and be better adapted to changes in temperature and nitrate

level than other species (Puche et al., 2018). Consequently, Confirmation of species

identification is reliant on genetic analysis and detailed observation of oospore

93

features with scanning electron microscopy (Casanova, 2005b), which was beyond

the scope of this study. Visual inspection in the field and subsequent keying out with

available dichotomous keys under light microscopy (Casanova, 2005b; Casanova,

2005a), left species identification unresolved to the genus level (Chara sp.), although

this identification exercise and comparison of samples collected between pools gave

confidence that the experiment was conducted on a single species alone.

The naturally nutrient depauperate conditions of the water column in these pools

would be expected to favour charophyte production over phytoplankton. Whilst

available N and P in both the water column and sediments are low, rates of

production by the charophytes in Coondiner Creek were two orders of magnitude

higher than phytoplankton per unit mass, and even higher when calculated by

biomass. Charophytes overall contribute a much larger proportion than

phytoplankton towards the total primary production within this creek system.

Christensen et al. (2013) reported similar strongly contrasting rates of production

between charophyte beds and phytoplankton in an oligotrophic shallow pond in

Sweden. Submerged macrophyte production was also ten-fold higher than

phytoplankton based on diel dissolved oxygen measurements made in tropical

floodplain wetlands (Adame et al. (2017). The rate of production by charophytes in

the present study is similar to rates measured for other Chara species under optimal

light and temperature (1.5 - 2.6 mg C g-1

DW h-1

) by Vieira and Necchi (2003), and

under natural conditions in Sweden at 0.39 - 4.15 mg C g-1

DW h-1

(Christensen et

al., 2013). Interestingly, these charophyte rates of productivity are also similar to

global average uptake rates (2.6 mg C g−1

DW h−1

) for seagrass communities

(Duarte et al., 2010), while a seagrass species occupying coastal waters and

94

measured in the cooler seasons produces at approximately half the rate (0.93 mg C g-

1 DW h

-1; Nayar et al. (2009). These findings suggest that charophytes may be

significant carbon sinks in IRES. Hence, any reduction in the suitability of pools to

charophyte growth – such as increased phytoplankton biomass and subsequent

change in light environment – may have ramifications for the carbon budgets of

IRES.

Charophyte beds contributed to significant shifts in diel dissolved oxygen levels in

the overlying water in these pools. In this study, very large ranges in dissolved

oxygen concentrations (from 1.84 - 14.94 mg L-1

) were recorded in pools that all had

abundant charophytes. Comparatively, Andersen et al. (2017) recorded dissolved

oxygen concentrations from 3.85 - 15.25 mg L-1

at the surface of charophyte

dominated shallow Swedish lakes. Under these supersaturated conditions, shading

under charophyte beds maintains a relatively anoxic layer at the sediment-water

interface during daylight hours (Andersen et al., 2017). This vertical stratification

may benefit heterotrophic respiration of benthic biofilms, hence enhancing the daily

differences in dissolved oxygen. The dense charophyte beds can carry a significant

epiphytic algae load that is likely to also contribute substantially to pool metabolism.

For comparison, a study from wetlands in tropical Northern Australia measured

epiphyton productivity in the range of 0.57 - 1.82 mg C g-1

DW h-1

depending on

morphology of the macrophyte to which they are attached (Adame et al., 2017).

Consequently, we would expect overall production occurring in charophyte beds (i.e.

including epiphytic algae) to be higher than charophytes alone as measured in this

study.

95

In this study overall metabolic responses between pools was highly variable,

although no difference between pools of contrasting hydrology was found. This

result is somewhat surprising given that previous studies have demonstrated that

connectivity to alluvial water can have a profound effect on organic matter and

nutrient fluxes (Fellman et al., 2011; Siebers et al., 2016). The results did indicate

high production and low production pools that weren’t related to pool hydrology.

Whilst hyporheic connectivity between surface water and alluvial groundwater

somewhat governs the hydrochemical nature of each pool, a larger effect on

phytoplankton and charophyte production is likely due to variations in temperature

as experiment bottles in all pools were light-saturated for the incubations. As

mentioned earlier, the variation in temperature among pools and possibly different

species involved, may partially explain why pool hydrology does not show an effect

on phytoplankton and charophyte production (Hill et al., 2009).

Measuring autotrophic response to nutrients in remote aquatic systems such as those

examined here remains challenging. While reasonably routine approaches to

measuring production were applied here, measurable growth responses to nutrient

additions were difficult. Nevertheless, tissue analysis and coupling of ambient

measurements versus more controlled experimentation and tissue stoichiometry

revealed N to be limiting in at least some pools. The findings also "hint" that

increased P in the water column may inhibit both charophyte and phytoplankton

short-term production in these pools.

There are a few factors that may also explain the lack of significant response to

nutrient additions in the incubation experiment. First, with hindsight the duration of

96

the experiments in this study (4 - 6 h) may have been insufficient for phytoplankton

and charophyte to utilise the excess nutrients immediately i.e. there is a lag between

nutrient uptake and metabolic or growth response. Metabolically, it has been shown

that when adding nutrients to extreme oligotrophic systems it often takes time for the

system to adjust before nutrient uptake can occur (Healey, 1979). ATP production in

photosystem-I takes precedence over carbon assimilation in the Calvin cycle.

Another explanation is that the response change in production is smaller than the

variability between sampling units. Both of these problems (lag-time due to initial

priming requirement, small response) could potentially be assessed by extending the

duration of the experiments to multiple days (Wehr, 1989; Elser et al., 2009),

although day-night cycles will presumably introduce other sources of variation to the

experiment.

Overall, this study demonstrates that charophytes are an important group when it

comes to net ecosystem production. Whilst the results of the main aims of this study

are inconclusive, they are largely consistent with other recent studies of charophytes.

Further investigation into nutrient response of species assemblages over longer

timeframe are warranted to further understand the complex roles of both N and P,

along with hyporheic exchange with alluvial groundwater in regulating production in

these pools.

97

5. CHEMOTAXONOMIC RESPONSES OF AUTOTROPHIC

PERIPHYTON COMMUNITIES TO NUTRIENT ADDITIONS IN

AN INTERMITTENT STREAM

5.1. Introduction

Periphyton communities are attached to surfaces in aquatic systems such as

sediment, rocks, wood or macrophytes and consist of a complex mixture of

autotrophic organisms such as green algae and cyanobacteria, as well as

heterotrophic microorganisms (Wetzel, 2001; Larned, 2010). The biomass and

taxonomic composition of periphyton communities are primarily shaped by the

energy inputs and nutrient status of the water body in which they reside (Townsend

et al., 2012), and are thus highly responsive to nutrient enrichment (Fairchild et al.,

1985; Tank & Dodds, 2003; Tank et al., 2017). Assessment of the periphyton algae

community can provide insight into changes in environmental conditions, and has

been used to indicate the onset of eutrophication (Gaiser et al., 2004) as well as

understand periphyton responses to change in hydrology (Townsend et al., 2012;

Sabater et al., 2016), light environment (Hill et al., 2010; Rier et al., 2014), and

dissolved gasses (Brown et al., 2017). Aquatic macrophytes (Chapter 4) and the

periphyton can contribute more autochthonous biomass than water column

phytoplankton in oligotrophic systems, as whilst the overlying waters are generally

impoverished in available nutrients, the sediment surface is a hotspot for nutrient

exchange (McClain et al., 2003). Periphyton has long been recognised for its

important functional role in the retention of nutrients in aquatic ecosystems,

especially of phosphorus (Reddy et al., 1999; Dodds, 2003; Scinto & Reddy, 2003).

98

Retention of nutrients by periphyton in shallow freshwater systems is further

enhanced by the settling of nutrient-bearing particles, along with efficient uptake and

recycling of nutrients between the autotrophic and heterotrophic component of

periphyton (Dodds, 2003). Consequently, we might expect that periphyton would be

an important nutrient recycler, and basal food source, in shallow pools of dryland

streams. However, despite the ecological significance of periphyton in IRES (Sabater

et al., 2016), the potential changes in periphyton community structure due to shifts in

nutrient availability or hydrologic status of IRES remain largely unknown for much

of inland and especially north Australia.

Intermittent streams in arid regions often depend on groundwater for pools to persist

beyond flood-flow events (Chapters 1 and 2). Groundwater mixing and discharge

into these pools via the hyporheic zone during these inter-flood periods is thus

critical for maintaining stream productivity (Burrows et al., 2018) and determines

carbon and nutrient cycling in pools in IRES and helps maintain higher trophic levels

(Fellman et al., 2011; Siebers et al., 2016). Pools that are not supplemented by

groundwater will undergo evapo-concentration of solutes during prolonged drought

periods with no surface flows (Fellman et al., 2011; Siebers et al., 2016). This

difference in carbon and nutrient status among pools (and seasons) will likely alter

both the biomass and composition of autotrophic periphyton communities and in

particular may result in shifts in dominance of green algae versus cyanobacteria. For

example, recent studies have revealed that P-Fe co-limitation can strongly limit

nitrogen-fixing cyanobacteria in aquatic ecosystems even when P is abundant

(Larson et al., 2018). Further, greater taxa richness and biomass of N2-fixing

organisms have been observed under treatments of phosphorus-iron addition

99

compared to treatments with only phosphorus addition (Larson et al., 2015).

However, the relative responsiveness of different taxonomic groups to

hydrochemical changes in streams with iron-rich sediments, and in particular to

increased nitrogen and phosphorus availability (Chapter 4), is unknown.

Biochemical approaches such as pigment analysis (Tamm et al., 2015) and

metagenomics (Friesen et al., 2017; Bengtsson et al., 2018) are increasingly being

used to characterise the functional taxonomy of freshwater periphyton. Specifically,

chemotaxonomy based on algal accessory pigments via high performance liquid

chromatography (HPLC) is an alternative and straightforward method which is

compatible with algal nutrient limitation experiments (Dalton et al., 2015). Nutrient

diffusing substrate (NDS) experiments conventionally measures Chlorophyll a

pigment as a response variable (Tank et al., 2017). With HPLC the experiment can

expand to also measure the response of algal accessory pigments produced by certain

taxa, and is capable of detecting algal taxa whose physical features are not well

retained in preservative. Chemotaxonomic analysis was developed and is extensively

utilised in marine systems as a means to detect, characterise, and monitor

phytoplankton communities (Wright et al., 1991; Jeffrey et al., 1999), although this

method would be similarly capable for periphyton. Individual pigments of interest

may be isolated for analysis, or the relative abundance of algal taxonomic groups in

the periphyton can be estimated by factor analysis from the calculated pigment ratios

(CHEMTAX: Mackey et al., 1996).

This study investigated the chemotaxonomic response of autotrophic periphyton to

nutrient additions on nutrient diffusing substrate (NDS) and how this response varied

100

among pools of contrasting connectivity to the alluvial groundwater under field

conditions. First, chlorophyll a biomass was measured to identify the extent to which

N and P availability limit periphyton production. Given grazing on algae can

influence overall periphyton responses to nutrient additions (Jones et al., 2000;

Eckert & Carrick, 2014), I compared the biomass of periphyton on nutrient enriched

substrates between open and caged experiments. Second, photosynthetic and

accessory pigments of the periphyton were quantified and Chl a:pigment ratios were

used to determine how autotrophic periphyton composition changed in response to

nutrient additions and how this varied with hydrology. The expectation was that

autotrophic biomass would be limited by both N and P, and that individual nutrient

additions would favour particular periphyton groups leading to distinct changes to

community composition. I also expected that both community compositional

changes and biomass responses to nutrient additions would be most apparent in the

most hydrologically isolated pools. (Jones et al., 2000; Eckert & Carrick, 2014)

5.2. Methods

5.2.1. Site description

As described in previous Chapters, Coondiner Creek is an intermittent – and

extremely ephemeral – dryland stream situated in the Upper Fortescue river

catchment of the Pilbara region of northwest Australia (Chapter 3: Figure 3.1).

Persistent and ephemeral pools (although generally recurring annually in the same

position) are most concentrated within the semi-confined gorge section of the creek

(see Chapter 4). Duration of water retention after cessation of surface flow is based

on a combination of pool hydrology, aspect/position, and channel substrate. The

101

following experiments were conducted at Coondiner Creek during a dry-phase in

July-August 2016. At this point in time surface flows had ceased ~18 months prior

and the stream had retracted to a series of isolated pools along the main channel.

Pools were classed by their hydrology as described in previous Chapters as either

predominately ‘persistent’ or ‘ephemeral’, based on previous studies of pool-

alluvium connectivity within the study catchment (Fellman et al., 2011; Siebers et

al., 2016). Pool hydrologic status was nevertheless confirmed for this study from

water stable isotope ratios measured at the time of sampling (see earlier Chapters).

5.2.2. Nutrient limitation experiments

Nutrient diffusing substrate (NDS) were constructed from 70 mL polypropylene

containers (Sarstedt, Germany) with a glass fibre filter (Whatman GF/F) acting as a

growth surface (Fairchild et al., 1985; Tank & Dodds, 2003). A round opening (⌀42

mm) was cut from the cap to expose the growth substrate. The containers were filled

with 2 % agar solutions amended with 0.5 M NH4NO3 (‘N’ treatment), 0.5 M

KH2PO4 (‘P’ treatment), 0.5 M NH4NO3 + 0.5 M KH2PO4 (‘NP’ treatment), or

unamended (‘C’ control treatment). To test whether grazing by fish and

macroinvertebrates had a significant impact on the biomass and composition of algal

development on the NDS, the nutrient design was duplicated with matching NDS

covered in 5 mm HDPE mesh. Five replicates of each nutrient treatment and grazing

experiment were attached to wire racks with the growth surface face-up and

positioned in either ‘persistent’ or ‘ephemeral’ pools. Samplers were left in situ to

incubate for 28 days, during which time there was no surface flow. After 28 days, all

samples were retrieved and the glass fibre filters removed. Filters and attached

periphyton were immediately placed in 5 mL 90 % acetone (AR grade; Chem-

102

Supply, South Australia), wrapped in foil and refrigerated for transportation back to

the laboratory in Perth within 3 days. The pigment acetone extract was filtered

through a 0.22 µm nylon filter (Thermo Scientific) into 1 mL glass HPLC vials

(Waters, Milford MA). Vials were capped and placed in -80 °C freezer until time of

HPLC analysis. A further 2 mL of sample was filtered and diluted to 20 mL in 90%

acetone for fluorometric determination of chlorophyll a. Sample extracts were

measured on a Trilogy fluorometer (Turner designs, San Jose, CA) using the non-

acidification method (EPA 445.0: Arar & Collins, 1997).

5.2.3. HPLC Pigment analysis

A subset of NDS were selected from the nutrient limitation experiment for pigment

analysis via high performance liquid chromatography (HPLC). The 60 sample subset

consisted of three replicates of each nutrient treatment per site from the ‘grazed’

treatment. A mix of standards and reference materials was used to build up a

pigment library and to calibrate pigments extracted from the samples. A mixed

phytoplankton standard PPS-MIX-119 (DHI group, Denmark) was also injected

once for each ten samples in order to evaluate drift in retention time throughout the

experiment. Peak areas were calibrated against chlorophyll a reference standard

(DHI group, Denmark). Algal reference material extracted from pure cultures of

Dunaliella tertiolecta (chlorophyte), Tetraselmis suecica (chlorophyte), Chaetoceros

muelleri (bacillariophyte) and Tisochrysis lutea (haptophyte) were also run. Method

blanks of 90 % acetone were processed identically to samples and passed through the

entire extraction process.

103

Pigments were quantified on a Waters HPLC system (600 controller, 217

autosampler, Waters, Milford MA) with a reverse-phase C18 column (Spherisorb

ODS2, 250 mm x 4.6 mm, 5 µm particle retention). Our solvents and elution scheme

were modified from Tamm et al. (2015). Solvent A consisted of 80 % methanol: 20

% 0.5 M ammonium acetate (pH 7.2) (v/v). Solvent B consisted of 80 % methanol:

20 % acetone (v/v). The elution scheme consisted of solvents A and B initially in a

50:50 mixture, switched to 100 % solvent B at 30 min, then returned to the 50:50

mixture at 50 min. Column flow rate was set at 0.7 mL min-1

and column

temperature was set at 22 °C for the duration of the experiment. Peaks were detected

with a 996 photodiode array (PDA) detector with scanning range 310 to 750 nm at a

resolution of 1.2 nm. PDA peaks were integrated at a quantification wavelength of

450 nm. Eluent then flowed through a 470 scanning fluorescence detector

(excitation: 440 nm, emission detection: 660 nm). Chromatographs were processed

in Empower2 software. Peaks were identified by comparison with standard reference

material and documented peak retention times and absorbance characteristics;

elution order and peak shape (Wright et al., 1991; Tamm et al., 2015). Peaks were

then integrated and peak area obtained using the software.

5.2.4. Pool hydrochemistry

Water samples were collected at the beginning and end of the incubation period and

analysed for nutrients (TDN: total dissolved nitrogen, SRP: soluble reactive

phosphorus), carbon (DOC: dissolved organic carbon, DIC: dissolved inorganic

carbon, SUVA254: specific ultraviolet absorbance at 254 nm), and stable isotopes of

water (δ2H, δ

18O), and dissolved inorganic carbon (δ

13C-DIC). Nutrient and carbon

samples were filtered through a sterile syringe filter (Sartorius mini-sart 0.45µm).

104

δ13

CDIC, δ2H and δ

18O isotope samples were filtered through a sterile syringe filter

(PALL 0.2 µm Supor) into a glass vial ensuring all headspace was removed. Samples

were immediately refrigerated (4 °C) in the field for transport back to the laboratory

for analysis. DOC and TDN were measured simultaneously on a Shimadzu TOC-V

analyser coupled with a total nitrogen module (Shimadzu TNM-1). Ultraviolet

absorbance at 254 nm was measured on a UV-Vis spectrophotometer (Cary 50,

Varian Medical Systems, Inc. CA USA). Specific ultraviolet absorbance (SUVA254)

was calculated using absorbance at 254 nm and DOC concentration as an estimation

of dissolved aromatic carbon content (Weishaar et al., 2003). Nitrate/nitrite and

ammonia were measured on a continuous flow analyser (Technicon Auto-analyser

II). Soluble reactive phosphorus (SRP) was measured spectrophotometrically by the

modified ascorbic acid method (Murphy & Riley, 1962; Kuo, 1996).

Water isotope samples were measured on a cavity ring-down spectrometer (Piccaro,

Santa Clara, CA, USA) following the analytical method outlined in Skrzypek and

Ford (2014), as described in earlier Chapters. A non-steady- state model was selected

for all pools as pool volume decreased in all pools during the experiment. The

evaporative loss fraction of the pool volume (f) over the duration of the experiment

was then calculated for each pool following Skrzypek et al. (2015), which is based

on a revised Craig-Gordon model (Craig & Gordon, 1965). The stable isotope

composition of the moisture in the ambient air (δA) was calculated using the isotopic

composition of the most recent large precipitation event proceeding the sampling

period and slope of the local evaporation line (LEL).

105


Statistical procedures were conducted in R (R Core Team, 2017) and PRIMER 6 &

PERMANOVA+ (Primer-E Ltd. UK). Pigments were quantified into major algal

groups following the CHEMTAX method (Mackey et al., 1996) with the limSolve

package in R (Soetaert et al., 2009). Initial pigment ratios for determining algal

groups were sourced from freshwater studies in the literature (Schlüter et al., 2006;

Sarmento & Descy, 2008; Dalton et al., 2015; Tamm et al., 2015), and multiple starts

were performed to ensure model convergence. Differences between nutrient and

hydrology factors were assessed using a permutational multivariate analysis of

variance (PERMANOVA) model (Anderson, 2001) for the response variables of i)

chlorophyll a biomass (µg cm-2

), ii) accessory pigment biomass (µg cm-2

) for each

individual peak detected, and ii) proportional taxonomic groups derived from

CHEMTAX analysis of Chl a: Pigment ratios. Univariate PERMANOVA of

chlorophyll a biomass was performed on a euclidean distance matrix produced from

log(x + 1) transformed values. The three-factor model had a crossed design with

pool hydrology (random: ‘persistent’ vs ‘ephemeral’), grazing (fixed: ‘grazed’ vs

‘ungrazed’), and nutrient addition (fixed: ‘C’, ‘N’, ‘P’, ‘NP’) as factors. Multivariate

PERMANOVA of accessory pigment biomass and proportional taxonomic groups

were performed on a Bray-Curtis similarity matrix of log(x + 1) transformed values.

The two-factor mixed effects model was designed with pool hydrology (random) and

nutrient treatment (fixed) as the factors. Each PERMANOVA model was run for 999

permutations with Type I (sequential) sum of squares. We report permutation p-

values at a significance level of α = 0.05. Multivariate data were visualised using the

distance based linear model DistLM procedure to produce dbRDA plots. Two-way

ANOVA were performed to compare pool water nutrient and carbon concentrations

106

between pool hydrology (persistent vs ephemeral pools) and time of sampling (initial

and final).

5.3. Results

5.3.1. Pool nutrients and hydrologic characteristics

The proportion of pool volume evaporative loss (f) ranged from 0.02 to 0.04 for

‘persistent’ pools whilst f ranged from 0.21 to 0.24 for ‘ephemeral’ pools, showing

that overall volumes of persistent pools over the 28 day experimental period

remained relatively constant whereas ephemeral pools lost up to one quarter of their

volume over the same period (Table 5.1). TDN ranged between 0.07 - 0.18 mg L-1

and the concentration was significantly higher in ephemeral pools than persistent

pools (ANOVA: F(1,26) = 8.01, p = 0.009), whilst there was no significant change in

TDN concentration over the course of the experiment. SRP ranged between 1 - 2 µg

L-1

with similar concentrations across both persistent and ephemeral pools, and did

not significantly change over the experimental period (Table 5.1). DOC ranged

between 1.30 – 3.59 mg L-1

and the concentration was significantly higher in

ephemeral pools than persistent pools (ANOVA: F(1,26) = 20.91, p < 0.001), whilst

there was no significant change in DOC concentration over the course of the

experiment.

5.3.2. Periphyton biomass response to nutrient additions

Chlorophyll a biomass ranged from 0.4 to 38.5 µg cm-2

across all treatments at the

end of the 28 day NDS experiment (Figure 5.1). Grazing exclusion did not

significantly affect chlorophyll a biomass (Pseudo-F = 1.61, P (perm) = 0.139).

107

There was also no significant interaction between grazing and nutrient treatments

(Pseudo-F = 0.844, p (perm) = 0.572). The three factor PERMANOVA model

showed that there was a significant difference in how the periphyton biomass

responded to nutrient availability between ‘persistent’ and ‘ephemeral’ pools

(Pseudo-F = 20.338, P (perm) = 0.001), with a significant interaction between

hydrology and nutrient treatment (Pseudo-F = 9.7, P (perm) = 0.001). In both

persistent and ephemeral pools, simultaneous N and P additions increased algal

biomass by more than three-fold compared to the control (Figure 5.1). Biomass also

more than doubled in ‘persistent’ pools in response to "N" alone (Figure 5.1). All

other treatments combinations showed no significant effect on periphyton biomass.

* Results of two-way ANOVA of log (x+1) transformed Chl-a data and its interpretation following

the methodology from Tank & Dodds (2003) are presented in Appendix 2 for reference.

PERMANOVA and the two-way ANOVA yielded similar findings.

108

Figure 5.1 Periphyton chlorophyll a response to nutrient additions in ‘persistent’ and

‘ephemeral’ pools. Nutrients added to substrates were nitrogen (N) as NH4NO3, phosphorus

(P) as KH2PO4, and nitrogen + phosphorus (NP). The control (C) received no nutrient

additions. The experiment was duplicated with ‘grazed’ and ‘ungrazed’ NDS treatments.

109

Tab

le 5

.1 C

har

acte

rist

ics

of

stu

dy p

oo

ls a

long C

oondin

er C

reek

at

init

ial

and f

inal

per

iod o

f per

iphyto

n i

ncu

bat

ion

. T

ota

l d

isso

lved

nit

rogen

(TD

N),

solu

ble

rea

ctiv

e p

ho

sph

oru

s (S

RP

), d

isso

lved

org

anic

car

bon (

DO

C),

spec

ific

abso

rban

ce a

t 2

54n

m (

SU

VA

254),

dis

solv

ed i

norg

anic

carb

on (

DIC

), s

table

iso

topes

of

filt

ered

wat

er s

ample

s (δ

13C

DIC

, δ

2H

, an

d δ

18O

), a

nd p

ool

evap

ora

tive

loss

(f)

.

110

5.3.3. Chemotaxanomic response of autotrophic periphyton

The identity of 21 chlorophyll and accessory pigments were determined in the

periphyton samples collected from the different substrates (Table 5.2). Good peak

separation was achieved for all the main pigments of interest with the exception of

diatoxanthin, which eluted within the broad peak base formed when lutein was

present in high concentrations (Figure 5.2a). Hence, we excluded diatoxanthin from

further analysis. Typical raw chromatograms from the sample set are shown here for

illustrative purposes (Figure 5.2b).

Table 5.2 Peak identification table of pigments identified in mixed

standard and periphyton samples.

111

Changes to community structure were evident in shifts in the biomass of accessory

pigments such as Fucoxanthin, Peridinin, and Lutein. The periphyton community

structure responded to nutrient availability differently between pools of ‘persistent’

and ‘ephemeral’ hydrology (Pseudo-F = 7.55, P (perm) = 0.001) (Table 5.3a). The

variation in response among pools and between hydrology to different nutrients can

be observed on axis-1 of the dbRDA plot, with more negative values corresponding

to higher biomass (Figure 5.3a). Axis-1 explains 63.7 % of the variation in the fitted

model. Many of the major pigments also align with axis-1 and are highly correlated

(e.g. Chl-a, Chl-b, Lutein, Fucoxanthin). Axis-2 accounts for 16.9 % of the variation

of the fitted model and distinguishes ‘ephemeral’ pools from, ‘persistent’ pools by a

greater response to the P treatment. This axis illustrates an increase in peridinin and

diadinoxanthin pigments, which are indicative of a higher proportion of

dinoflagellates in ‘ephemeral’ pools, especially with P treatment.

112

Periphyton pigment compositions are considered representative of changes in major

taxonomic groups. Based on the CHEMTAX approach, communities colonising the

control NDS (no nutrient addition) were estimated to consist of 60 % diatoms, 13 -20

% chlorophytes, 7-12 % euglenophytes, 9 % cyanobacteria, and 7 % dinoflagellates

Figure 5.2 HPLC chromatograms showing a) standard pigment mix, peak numbers

correspond with those in Table 5.2, and b) a typical HPLC chromatogram from a persistent

pool showing control (black), nitrogen (red), phosphorus (blue), nitrogen + phosphorus

(green). Absorbance was measured at 450 nm

113

in both ephemeral and persistent pools (Figure 5.4). However, nutrient treatments

differed in their periphyton community structure between pool hydrology (Pseudo-F

= 2.55, P (perm) = 0.030). Relative to the control, P treatments in ‘persistent’ pools

had a decrease in the proportion of diatoms (P: 37 %, NP: 45 %; F = 10.86, p =

0.002) and euglenophytes (P: 3 %, NP: 1%; F = 39.09, p < 0.001), and an increase in

the proportion of chlorophytes (P: 37 %, NP: 38 %; F = 28.98, p < 0.001). There was

no change in the proportion of dinoflagellates. In contrast, P treatments in

‘ephemeral’ pools showed decreased proportions of diatoms (P: 18 %, NP: 22 %; F

= 32.95, p < 0.001), and an increase in dinoflagellates (P: 44 %, NP: 36 %; F =

13.92, p = 0.001) relative to the control, but no change in the proportion of

chlorophytes (Figure 5.4). The community structure of N treatments in ‘persistent’

pools was similar to the control, although the proportion of cyanobacteria was

reduced (F = 12.54, p = 0.001). In contrast, in ‘ephemeral’ pools, the proportion of

diatoms (N: 36 %, NP: 45 %), dinoflagellates (N: 25 %, NP; 36 %), and

euglenophytes (N: 14 %) all increased when N was added either alone or with P.

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The periphyton community responded significantly differently between pools of

‘persistent’ and ‘ephemeral’ hydrology based on CHEMTAX analysis (Pseudo-F =

14.767, P(perm) = 0.001; Table 5.3b). For both ‘persistent’ and ‘ephemeral’ pools

we observed a shift away from diatom dominated periphyton community when

Figure 5.3 Multidimensional dbRDA plots of pigments extracts from the periphyton NDS

experiment; a) pigment biomass (µg cm-2), and b) estimates of algal group proportions by

CHEMTAX analysis. Results are based on a Bray-Curtis similarity matrix of log(x + 1)

transformed samples (n = 60).

Table 5.3 Factorial two-way mixed effects PERMANOVA of a) periphyton pigment

biomass (µg cm-2

), and b) estimates of algal group contributions from CHEMTAX analysis

of Chl a: Pigment ratios. Pool hydrology and nutrient treatment are included as factors.

Significant P-values are indicated in bold.

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nutrients were non-limiting. For ‘persistent’ pools I observed a shift from diatoms to

chlorophyta, whilst in ‘ephemeral’ pools there was a shift towards a dinoflagellate

dominated periphyton community. These results are graphically illustrated on the

dbRDA plot with diatoms, chlorophytes, and dinoflagellates separated out strongly

(Figure 5.3b). Axis-1 aligns with communities being chlorophyte or diatom

dominated and explained 58.4 % of the variation of the fitted model. Axis-2 of

dbRDA plot aligns with an increase in the proportion of dinoflagellates and

explained 35 % of the variation of the fitted model.

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Figure 5.4 Estimates of algal group contributions to periphyton community structure

calculated from Monte Carlo perturbations of CHEMTAX analysis. Nutrients added to

substrates were nitrogen (N) as NH4NO3, phosphorus (P) as KH2PO4, and nitrogen +

phosphorus (NP). The control (C) received no nutrient additions. Mean proportion of each

group per nutrient and hydrology treatment is shown with standard error (n = 3).

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5.4. Discussion

This study improves our understanding of freshwater algae communities in

northwest Australia (Masini, 1988, 1989; McIntyre, 2009), and is the first to directly

test nutrient limitation on periphyton communities in this region. My results

demonstrate that autotrophic periphyton productivity in Coondiner Creek is nitrogen

and phosphorus co-limited regardless of whether pools were ephemeral or persistent

in hydrology. I observed subtle differences in both chlorophyll a biomass response

and shifts in community structure among pools of differing hydrology. Taxonomic

responses to nutrients varied considerably with hydrology: in general ‘persistent’

pools shifted towards a chlorophyta-dominated community, while ‘ephemeral’ pools

shifted towards a dinoflagellate-dominated community.

Periphyton production (chlorophyll a biomass) was co-limited by N and P. The

combination effect of adding both N and P causes a synergistic response in

autotrophic production. The responses observed in the pools investigated in the

present study are consistent with widespread co-limitation by N and P that has been

observed across a broad range of aquatic systems (Francoeur, 2001; Elser et al.,

2007). In Coondiner Creek, an (as yet) relatively pristine environment, periphyton

communities are most probably highly adapted to scavenging nutrients that may only

be episodically available via increased uptake efficiencies and nutrient recycling

between autotrophic and heterotrophic components (e.g. Scinto & Reddy, 2003).

Phosphorus was the primary nutrient responsible for shifts in periphyton community

structure. Other periphyton limitation studies have indicated that while phosphorus

addition may promote algal biomass, it does not generally result in a change in

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periphyton community structure in freshwater streams (Dalton et al., 2015; DeNicola

& Lellock, 2015; Vizza et al., 2018). Within Australia, Townsend et al. (2012) found

no change in autotrophic periphyton community structure in a tropical stream

between control and nutrient additions. However, nutrient concentrations in the

Townsend et al. (2012) study were only marginally above ambient conditions. In an

alpine stream, pulses of P were found to result in a reduction in periphyton species

diversity (Davies & Bothwell, 2012). Hence, the source and duration of phosphorus

enrichment to periphyton, rather than the P concentration per se may be more

controlling of community structure.

This in situ study also showed little impact of grazing on periphyton biomass over 28

days, which was a surprising finding. It was assumed prior to deploying the NDS

experiment that grazing by fish and macroinvertebrates would have a negative effect

on the periphyton biomass (e.g. Hillebrand & Kahlert, 2001; Hill et al., 2010). We

observed fish schools within all pools when selecting sites for this study and as

isolated pools act as refuge for native fish populations between flow events (Morgan

& Gill, 2004; Beesley & Prince, 2010; Lostrom et al., 2015), predation pressure by

fish such as Rainbowfish (Melanotaenia australis) would presumably increase as the

pools contract, thus reducing grazing effects of macroinvertebrates on periphyton.

Alternatively, the mesh may have protected small (< 5 mm) grazers from fish

predation. Shading is well known to be a limiting factor in periphyton production

(Von Schiller et al., 2007; Hill et al., 2009; Guo et al., 2016). The mesh used in this

study blocked ~ 5 % light onto the GF/F. Hence, the shading effect would be

minimal and possibly insignificant.

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This study clearly demonstrates that chemotaxonomic analysis is an effective method

for assessing changes in periphyton community structure. Chemotaxonomic analysis

may be a relatively inexpensive and straightforward approach for monitoring

responses to environmental change e.g. altered flows or nitrate inputs from mining

discharge (Dogramaci et al., 2015; Degnan et al., 2016), increased concentrations

owing to reduced flows (Siebers et al., 2016; Bestland et al., 2017), increased inputs

from dust deposition from fertilizer applications to surrounding catchments and N

and P from cattle (McDowell & Stewart, 2005; Pettit et al., 2012). Overall, these

results demonstrate that periphyton biomass in Pilbara streams is sensitive to both

nitrogen and phosphorus inputs. Surface and groundwater runoff of nitrogen, along

with atmospheric deposition are increasing in northwest Australia due to industrial

activities (fertilizer production, disturbance and airborne dust from resource

extraction), and agricultural (nutrient supplemented irrigation schemes, rangeland

grazing) sources. Hence, these new N sources have the potential to increase rates of

periphyton production in Pilbara streams, which in turn may affect higher trophic

orders.

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6. GENERAL DISCUSSION

6.1. Overview

The findings of this thesis significantly increase knowledge of the key processes that

drive productivity and metabolism of pools in intermittent rivers and ephemeral

streams (IRES) of hot and arid environments. The results of this research also

demonstrate that IRES in the Pilbara region can be both N and P co-limited; however

the different sensitivities and /or responsiveness among producers to nutrient

additions (i.e. charophytes versus periphyton) also illustrate the complexities of

predicting ecosystem responses to anthropogenic pressures. In these Pilbara streams,

I have shown that between-pool variation in groundwater connectivity also interacts

with nutrient availability such that persistent pools may be more sensitive to excess

N compared to more ephemeral pools. In this final chapter, I discuss the major

findings of the research presented in this thesis within the broader context of current

understanding of how IRES function in hot and arid landscapes. I also consider the

implications of the research findings for the future management of Pilbara streams,

particularly where faced with altered hydrology and changing catchment land use.

6.2. Alluvial groundwater connectivity influences stream biogeochemistry

and metabolism

A strong underlying theme throughout this thesis has been developing an

understanding of the importance of alluvial groundwater connectivity to stream pool

biogeochemistry. Vertical hydrological exchange between the surface water and the

hyporheic zone mediates transport of products – such as mineralised or desorbed

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phosphorus (Chapter 3) – from biogeochemical activities within the sediments

(Boulton et al., 2010). While hyporheic exchange and the biogeochemical processes

within the sediments are known to occur across multiple hierarchical spatial scales,

there have been remarkably few studies to date that have investigated these

interactions in arid environments. The research presented here seeks to address that

knowledge gap. In Chapter 2, I demonstrate that stream pools at both reach and

catchment scales are strongly shaped by hydrological connectivity. The responses of

the biotic components of pools to nutrient addition are also mediated by the degree

of groundwater connectivity (Chapter 5). Persistent pools also showed much greater

evidence of N limitation than more ephemeral pools. While I observed some

catchment-wide trends, such as increasing solute concentration with evaporation,

none of the typical longitudinal patterns described by the River Continuum Concept

(Vannote et al., 1980) were observed. This suggests that stream models developed

for describing temperate systems may not be suitable for explaining how intermittent

dryland streams function. It is likely that in the Fortescue River catchment local

environmental factors such as hydrology and nutrient cycling generally dominate

over catchment-scale processes.

In this semi-arid landscape stream pool persistence is prolonged by connectivity to

alluvial water. This can lead to higher in-stream productivity, as productivity is

directly dependant on the extent and duration of surface water availability (Bunn et

al., 2006a; Larned et al., 2010; Leigh et al., 2010). However, pool-scale metabolism

measured during this study could not be predicted by connectivity to the alluvial

water (Chapter 4). Other factors such as daily irradiance and tight nutrient cycling

more likely play a substantial role in IRES pool metabolism rates. As these streams

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cut their way through the Hamersley Ranges, the overall complex combination of

alluvial connectivity, riparian cover, and topography (cliff aspect/shading) governs

the diel timing and amount of irradiance reaching these pools. Additionally, I

expected to see hyporheic processes, such as mixing and upwelling of nitrate from

groundwater, enhance nitrogen available in persistent pools (Grimm, 1992; Boulton

et al., 2010), whereas I presumed this would be less important in in ephemeral pools

where hyporheic exchange is much less or non-existent. However, in this study

productivity was not enhanced in persistent groundwater-connected pools. One

possible explanation for this is that much of the nitrogen made available may be

directed to rooted macrophytes and emergent vegetation that acts as a nutrient sink

rather than to water column productivity. Thus, enhanced production from these

nitrate inputs may not be measureable using changes in aquatic oxygen

concentrations as emergent macrophytes directly exchange oxygen with the

atmosphere rather than through the water column. It is of interest that metabolism

within these pools was net autotrophic irrespective of the degree of alluvial

connectivity. This result is contrary to the general assumption that once nutrients are

exhausted and pools become isolated that metabolism would be net heterotrophic

due to the increased rates of organic matter breakdown (Corti et al., 2011; Datry et

al., 2018b).

There is much diversity in biogeochemical processes within streams and across

catchments leading to high diversity in biotic character, where no two stream pools

are the same. Hence, models which incorporate this high spatial and temporal

variability across a hydrological discontinuum are necessary to frame ecological

processes in intermittent stream systems. The Punctuated Biogeochemical Reactor

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(PBR) model (Larned et al., 2010) may be a more appropriate model for describing

stream processes within the Fortescue River catchment. This model is based on

material processing rates, and essentially incorporating flow variability to form an

extension of the nutrient spiralling concept (Newbold et al., 1982). The PBR

proposes that within intermittently flowing rivers material processing is rapid during

inundation periods and processing efficiency increases the further the material is

transported downstream. Whilst the PBR model is restricted to simplified contrasts

between wet and dry periods, IRES are generally much more dynamic, with

hydrological periods dominated by expansion, contraction, fragmentation and

desiccation. Hence, in IRES such as in the Pilbara – where isolated stream pools

persist, prolonging the fragmentation period – a significant contribution to material

processing also occurs. For example, DOM transitions to more labile forms over

periods of extended isolation of persistent pools (Fellman et al., 2011; Siebers, 2015;

von Schiller et al., 2015). Thus, these pools are potentially highly important for

‘priming’ of heterotrophic instream processes throughout the catchment when

subsequent flows occur (Guenet et al., 2010).

6.3. Sediment mineralogy constrains within-stream nutrient bioavailability

This study revealed that both nitrogen and phosphorus are strongly limiting at both

the catchment and local stream pool scales. Phosphorus availability in the water

column is constrained in Pilbara streams by i) sediment-P sorption, primarily with

clays and iron-rich minerals, and presumably ii) Calcium-phosphorus co-

precipitation or sorption where Ca is present either in solution or as calcium

carbonate in the sediment. Iron-rich stream sediments in Coondiner Creek acted as a

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strong abiotic sink for phosphorus, with adsorption capacities far exceeding

phosphorus inputs (Chapter 3). Calcium-phosphorus co-precipitation may be

especially important in regions where carbonate aquifers are the predominant water

source, such as seen in this study at Weeli Wolli Creek, where outcrops of Ca-

bearing calcareous deposits are widespread. Calcium- and/or magnesium- carbonates

also play a role in P sorption processes in sediments (Pant & Reddy, 2001). Pure

calcite has an exceptionally high P sorption capacity (Moharami & Jalali, 2013), and

Ca-P sorption is high in a temperate intermittent river where sediments are

dominated by calcium carbonate (Jalali & Peikam, 2013). Hence, for parts of the

Pilbara region where stream sediments are both iron- and calcium- rich we would

expect abiotic P limitation to be especially strong.

This study demonstrated that moderate DOM additions at similar concentrations to

what we see in evaporated pools reduced sediment adsorption capacity. Chapter 2

demonstrated that there was a relationship between how evaporated a pool is and the

DOC content. Hence, this evapo-concentration of DOC may be reducing the

adsorption capacity of sediments, and therefore making it more likely for

remineralised organic P and any new inorganic P entering the stream to remain in a

bioavailable form. This may partially explain why we see a second flush of

production occurring as these pools dry down. There was a preferential sorption of

high spectral slope DOM to sediments when P is absent or at low concentrations.

Hence, this may partially explain why the proportion of labile DOM increases over

periods of extended isolation (Fellman et al., 2011; Siebers, 2015; von Schiller et al.,

2015). Interactions between organic matter and nutrients are important in regulating

nutrient availability in streams (Coble et al., 2016; Porcal & Kopacek, 2018). Adding

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allochthonous organic matter to pools directly affects the P adsorption capacity of

sediments and release of Fe-bound phosphate from sediment sinks (Chapter 3).

Altered organic matter, nutrient loads, and/or hydrology due to land use change may

thus have significant but as yet poorly understood impacts on the ecological

functioning of intermittent streams.

The preferential desorption of DOM from sediment at elevated inorganic phosphorus

concentrations (Chapter 3) suggests a potential ‘shunting’ mechanism where

inorganic P is rapidly removed from the bioavailable pool when it come in contact

with these Fe-rich sediments (via adsorption), and releases loosely sorbed organic P

compounds. These organic P compounds may then be more readily utilised by

microbial communities beyond the sediments (e.g. epiphytic and pelagic

heterotrophic bacteria). The exact mechanisms are unclear, although may be an

interesting avenue to pursue in relation to understanding the dynamics of organic

phosphorus in streams (Baldwin, 2013). I undertook a pilot study to characterise

organic P in these Pilbara stream sediments in the early stages of this research

(Appendix I). However, organic P contents were low, and difficult to discern in these

sediments against baseline noise due to interference from iron. While not conclusive,

these initial studies identified orthophosphate monoesters (phytic acids),

orthophosphate diesters (nucleic acids) and pyrophosphates in a range of Pilbara

sediments. Further work pairing batch sorption isotherm experiments with 31

P-

nuclear magnetic resonance to observe changes in organic P associated with

sediments would help elucidate further interactions between DOM-P and

geochemistry in these systems.

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6.4. Complex responses of aquatic primary productivity to perturbations in

nutrient status in dryland streams

This study showed contradictory results regarding which nutrients were limiting

productivity and is an example of why assessing nutrient limitation at multiple scales

can provide useful insights into the operation of biogeochemical processes within the

system. The catchment-wide stoichiometric approach (Chapter 2) suggested that the

system was P-limited. However, the process-based nutrient limitation studies at the

pool-scale (Chapter 5) demonstrated N and P co-limitation of production, which is

common in freshwater systems globally (Elser et al., 2007). Substantial previous

work has shown that the terrestrial environment (e.g. soils, agricultural production,

and plant communities) of the Pilbara region is also generally depauperate in

phosphorus (Bentley et al., 1999; Islam & Adams, 2001; McIntyre et al., 2009a;

McIntyre et al., 2009b). Within streams, periphytic algae which have evolved in P-

limited systems may be especially effective at retaining P (e.g. Dodds, 2003).

Additionally, heterotrophic microbial uptake and remineralisation of iron-bound P

(otherwise considered biologically unavailable) may be important pathways for

delivering inorganic P to autotrophic organisms (Jaisi et al., 2011). Therefore

capturing additional local pool-scale processes may fill important research gaps to

fully understand nutrient limitation in these systems.

The timing and duration of nutrient limitation experiments may be highly relevant to

a more thorough understanding of biogeochemical processes in this system. An

immediate short-term response is unlikely as the algae in this system (especially

charophytes) are presumably highly adapted to living in an environment depauperate

in P. Longer incubation time allow for algal cells to respond to new conditions, also

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to flush/deplete internal stores of nutrients to levels where cells are forced to depend

on external inputs. Productivity and biogeochemical processes in these streams may

also vary over the wet-dry hydrological cycle (time-scales of ~seasons to years). It is

likely also that the system fluctuates between primarily N or P limitation through the

seasons and the streams natural hydrological phases (Reisinger et al., 2016). Whilst

characterisation and identification of biogeochemical processes during this ‘dry’ part

of the hydrological phase is important, we are missing the most energetic and

dynamic period within these systems. How nutrients and carbon are transported and

processed during flood-flow events has not been characterised in this study.

However, these first flows are important release of remineralised nutrients and

carbon upon rewetting of sediments (Baldwin & Mitchell, 2000). Consequently,

further investigations which capture the distinct hydrological phases that distinguish

IRES from perennial systems may provide further insights into stream nutrient

processes.

Multi-pigment analysis was used in this study to compliment the more traditional

chlorophyll a methods as a means to collect quantitative periphyton community

response data (Chapter 5). Pigment analysis is useful for characterising algal taxa

including when cryptic or difficult-to-preserve species are present. Another

promising approach consists of metagenomic sequencing of periphyton communities

as a method to simultaneously assess not just autotrophic algae, but also the

heterotrophic bacteria, fungi, and other microbes. Tools that further assess

community structural response may be a move towards novel and bespoke methods

for biomonitoring in IRES (Stubbington et al., 2018) although metagenomics

techniques such as eDNA require an increased understanding of how genetic

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material is processed within streams (Shogren et al., 2017). Early adopters of eDNA

techniques have mainly been interested in its potential for cataloguing species

(Carew et al., 2017) along with the detecting and monitoring of endangered or

ecologically relevant species (Thomsen & Willerslev, 2015). Metagenomics have

elsewhere been utilised for the characterisation of periphyton communities (Friesen

et al., 2017). However, a breakthrough in obtaining quantitative sequencing is

required to enable eDNA techniques to be suitable for quantitative characterisation

of periphyton communities.. These emerging techniques in ecological "'omics" are a

promising avenue for future studies of the links between community compositional

and functional changes in response to nutrient limitation studies and process

measurements (i.e. productivity rates). .

6.5. Implications from this research to understanding responses of stream

ecosystems in northwest Australia to changing land use and climate

There is increasing concern over the cumulative impacts of mining and agriculture to

the ‘health’ of intermittent streams in the Pilbara. These impacts are coupled with the

uncertainty of how climate change will affect the hydrology across the region. There

is potential for atmospheric and runoff related deposition of nitrogen and phosphorus

to increase throughout northwest Australia. This is due to increase in nitrogen release

from industrial (fertilizer production, disturbance and airborne dust from resource

extraction), and agricultural (large scale nutrient supplemented irrigation schemes,

rangeland grazing) sources. Mining has resulted in numerous stream diversions and

mine voids (pit lakes) impacting the active stream, parafluvial zones, and

groundwater aquifers (McCullough & Lund, 2006; Rojas et al., 2018). A number of

major streams draining the Hamersley Ranges are also being used for current and

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future mine wastewater discharge operations (Dogramaci et al., 2015; Cook et al.,

2016) (Figure 6.1). Hence, significant changes in groundwater/surface water

hydrology are increasingly widespread across the Pilbara due to both the direct and

cumulative impacts of resource extraction throughout the catchments. Agricultural

practices, such as unregulated cattle grazing in streams and riparian areas (Lyons,

2015), invasion of riparian zones by buffel grass (Cenchrus ciliaris) (Miller et al.,

2010; Marshall et al., 2012), and irrigated fodder crop production (Schelfhout &

Broad, 2015) all have possible impacts on delivery of allochthonous OM to streams.

Dense buffel grass can replace a sparse mixed native herb and grass understorey,

making riparian corridors much more susceptible to bushfire (Pettit & Naiman,

2007; Miller et al., 2010). Soil remineralisation due to fire alter nutrient and carbon

dynamics – in turn affecting stream metabolism through increased inputs – and also

increase the ‘flashiness’ of flows (Pettit & Naiman, 2007; Robson et al., 2018).

Furthermore, the increase in exotic grass biomass has a potential to shift the carbon

source of allochthonous OM from predominantly woody to more labile grassy

materials for aquatic food webs. It is unclear how sensitive IRES are to changes in

these carbon sources or increased catchment loads of nitrogen and phosphorus.

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While a region of extreme climate variability, the Pilbara region has showed

noticeable wettening in recent decades compared to previous centuries (Cullen &

Grierson, 2007; O'Donnell et al., 2015; Rouillard et al., 2015; Rouillard et al., 2016).

Future climate modelling scenarios predict increased air temperature, with the

delivery of rainfall projected to also change, with a reduction in frequency, but

increase in intensity, of tropical cyclones (Charles et al., 2015; Sudmeyer, 2016).

Importantly, potential evaporation is also projected to increase (Charles et al., 2015),

which will increase the rainfall deficit, and directly affect surface waters. Hence, the

Figure 6.1. Examples of the diversity of hydrologies and settings of streams in the central

Pilbara. Weeli Wolli Creek, a spring-fed creek in the Hamersley Ranges during a) dry

periods receiving minewater discharge, and b) moderate flood after a 30 mm rainfall event.

Note the significant increase in suspended sediments during flood events. c) and d) Typical

catchment vegetation in the Hamersley Ranges. Many hilltops and slopes have sparse

vegetation on highly weathered skeletal soils.

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extent of stream surface water throughout the region may be expected to contract and

become fragmented more rapidly after flow cessation – a key process in IRES.

Future climates may alter both the extent and duration of surface water, and the

evapo-concentration of nutrients, increasing nutrient retention in isolated stream

pools (McLaughlin, 2008). For streams where evaporation already plays a critical

role in shaping the differences between persistent (alluvially connected) and

ephemeral pools along these streams will increase. Spring-fed ‘persistent’ and the

arguably more vulnerable ‘ephemeral’ pools across the region both have unique

character and will require complementary approaches to enable each to be managed

sustainably. Understanding and managing impacts against a background of extreme

variability remains a challenge.

6.6. Conclusion

The findings of this thesis is particularly timely for intermittent rivers and streams in

the semi-arid areas of Australia in predicting the effects of impending development

on relatively undisturbed riverine landscapes. The information discussed in this

thesis provides important insights into the critical ecological processes in a semi-arid

river catchment. The functioning of intermittent rivers such the one studied here is

characterized by extremely high complexity of ecological processes and highlights

the importance of longitudinal, lateral and vertical linkages of water to the stream.

This study has been the first in the region to mechanistically investigate a key abiotic

process behind nutrient limitation. Also, this study utilised novel but relatively

straightforward approaches to directly measure response of primary producers to

nutrient additions in remote and logistically challenging field conditions. I found that

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these intermittent hot arid streams are highly dynamic and not well-described by

current stream models. Productivity and ecological community structure is

influenced by several abiotic factors including sediment mineralogical interactions,

groundwater connectivity, and catchment hydrological processes, as well as biotic

factors such as large inputs of allochthonous carbon from riparian vegetation.

Although this thesis has provided some key details of nutrient limitation and the role

of carbon in P availability, there is still limited knowledge of how fluxes of different

forms of carbon, especially terrestrial carbon, influence these stream processes. We

still have only limited knowledge of how nutrient transfers and cycling drive primary

production within these systems where P is tightly bound to sediments and therefore

unavailable. In addition, our knowledge is limited on the mechanisms involved in the

tight cycling of nutrients between the sediments, water and plants and the role of

microorganisms. There is still insufficient knowledge of how this systems work to

predict ecological responses to either human-generated or climate-related changes.

Our limited understanding of the linkages between biogeochemistry and food web

processes in IRES is a key knowledge gap. Studies are required that merge terrestrial

and aquatic ecology with biogeochemistry and hydrology to address this.

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7. APPENDIX 1 - PILOT STUDY INVESTIGATING THE

SUITABILITY OF 31

P-NMR FOR THE CHARACTERISATION OF

ORGANIC PHOSPHORUS IN IRON-RICH PILBARA STREAM

SEDIMENTS

7.1. Methods

Sediment samples were collected from pools along an intermittent stream

(Coondiner creek) in the Pilbara in October 2013. Sediment was sampled by

collecting the top 5cm of sediment using a corer (diameter = 60 mm). The sediment

core was sieved through 2 mm mesh and the retained <2 mm fraction was dried (60

°C) and ground in a ball mill to a fine powder prior to sample extraction. A sub-

sample of three sediments was selected from the set for this pilot study.

7.1.1. Sample pre-treatment

Samples were subjected to standard NaOH + EDTA extraction (treatment 1)

(Vestergren et al., 2012; Özkundakci et al., 2013) and two sodium dithionite

extraction treatments based on methods proposed in the literature (Reitzel et al.,

2012; Zhang et al., 2013) to reduce iron (Fe) interference in subsequent 31

P-nmr

experiments. Treatment 1 - Sediments were extracted in 0.25 M NaOH + 0.05 M

EDTA at a 1:4 w/v sediment:solvent ratio. Samples were shaken for 16 hr in a rotary

mixer and then centrifuged (2000 rpm for 30 min) and the supernatant retained.

Treatment 2 – Sediment was first extracted in 0.11M NaHCO3 + 0.11M Na2S2O4 at a

1:4 w/v sediment:solvent ratio. Samples were shaken for 2 h in a rotary mixer and

then centrifuged (2000 rpm for 30 min) and the supernatant discarded. Retained

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solids were then extracted as per treatment 1. Treatment 3 – steps taken as per

treatment 2. 2% 0.11 M NaHCO3 + 0.11 M Na2S2O4 was then added to supernatant

and centrifuged (2000 rpm for 5 min). A 10 mL aliquot of each sample was retained

for ICP analysis and inorganic P by colorimetry. The remaining extract was

lyophilised for 31

P-nmr experiments.

7.1.2. Sediment chemistry

Elemental analysis of sediment extracts was conducted on a PerkinElmer optima

5300DV inductively coupled plasma optical emission spectrometer (ICP-OES).

Elements measured were Al, Ca, Fe, Mg, Mn, P, and S. Inorganic P (Pi) was

measured by the molybdenum-blue ascorbic acid method (Murphy & Riley, 1962).

Organic P (Po) was calculated as the difference between total P determined by ICP-

OES and Pi.

7.1.3. 31P-nmr experiment

Solution 31

P-nmr spectra were acquired at 298 °K on a Bruker AV 500 spectrometer

(Bruker, Germany) at a 31

P frequency of 202.5 MHz. Sample extract was added to a

10 mm glass nmr tube. We used a 90 ° pulse of 41.0 µs, an acquisition time of 1.0 s.

5120 scans were acquired. Chemical shift was referenced to orthophosphoric acid at

5.556 ppm. The spectra presented have a line broadening of 20 Hz and was not

decoupled.

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T

ab

le 7

.1 C

om

posi

tion o

f se

dim

ent

extr

acts

det

erm

ined

by I

CP

-OE

S a

nd

co

louri

met

ry.

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7.2. Results

7.2.1. Sediment and extract chemistry

Extract total P measured by ICP-OES correlated with inorganic P measured by

colourimetry (Figure 7.1, R2

= 0.98). This suggests that the majority of P present is

in the inorganic form. Total P ranged from 14 to 100 mg/kg. More P was extracted

with the dithionite treatments (T2 and T3) for sediments HD4002B and WINA

although there was little difference for sediment HD4007A. Iron (Fe) ranged from 2

to 2938 mg/kg with less Fe in the extract for the NaOH + EDTA treatment (T1) than

the two dithionite treatments (T2 and T3). There was an inverse relationship present

between Ca and P, and possibly between Ca and Al.

Figure 7.1 a) Comparison between total P measured by ICP-OES and inorganic P measured

by colourimetric detection. Dashed line indicates 1:1 relationship, solid line indicates linear

regression (R2 = 0.98), b) Comparison between total P and Fe, c) P and Ca (note: log10

scale on y-axis), and d) Al and Ca (note: log10 scale on y-axis) measured by ICP-OES.

Treatments T1: NaOH+EDTA, T2: dithionite before, T3: dithionite before and after, see

methods for detail.

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7.2.2. 31P-nmr spectra

31P-nmr detected 4 forms of organic phosphate within sediments from Coondiner

creek (Figure 7.2). Peaks were detected at 4 ppm for orthophosphate monoesters

(phytic acid), -1 ppm for orthophosphate diesters (e.g. DNA), -5 ppm for

pyrophosphates, and -23 ppm for polyphosphates. The large peak at 6 ppm

corresponds to inorganic orthophosphate.

7.2.3. Comments on method suitability

Treatment 1 (NaOH + EDTA) was determined to be most suitable for 31P-nmr

analysis. Both treatment 2 and 3 increased the amount of iron in the final sediment

Figure 7.2 Solution 31

P-nmr spectra of NaOH-EDTA soil extract from Window pool (WINA-

t1), Coondiner creek. Prepared on a) Brucker 500 in a 10 mm tube, and b) Brucker 600 in a 5

mm tube. The vertical scale has been exaggerated 10x on the upper trace to delineate

individual peaks.

138

extract. Treatments 2 and 3 (Dithionite extraction steps) were unsuitable as it sends

Fe to the liquid phase causing Fe to go into the extractant rather than being

precipitated out of it. Possibly need to use a higher molar concentration of dithionite

to account for the very high concentration of Fe in sediments. The limit here is that

the reaction requires enough S to react with Fe. The dithionite used in these

extractions was of questionable quality as the bottle had been opened in 2007.

The sample extracts had a lot of iron oxides precipitating out. The reaction only

works if kept in a reducing state as oxidisation causes the Fe to precipitate as iron

oxide. Future studies should take note of redox potential throughout the extraction

process.

139

8. APPENDIX 2 – TWO-WAY ANOVA

Table 8.1 Results of two-way ANOVA of log(x+1) transformed Chlorophyll-a data and its

interpretation following the methodology from Tank & Dodds (2003).

Source df SS MS F-value P-value

a) Persistent pools (grazed)

Nitrogen 1 23.308 23.308 168.69 < 0.001

Phosphorus 1 4.117 4.117 29.8 < 0.001

Nitrogen:Phosphorus 1 3.457 3.457 25.02 < 0.001

Residuals 56 7.737 0.138

b) Ephemeral pools (grazed)

Nitrogen 1 5.189 5.189 56.08 < 0.001

Phosphorus 1 6.89 6.89 74.46 < 0.001


Residuals 36 3.331 0.093

c) Persistent pools (ungrazed)

Nitrogen 1 13.87 13.87 82.75 < 0.001

Phosphorus 1 2.065 2.065 12.32 < 0.001


Residuals 56 9.387 0.168

d) Ephemeral pools (ungrazed)

Nitrogen 1 4.933 4.933 21.317 < 0.001

Phosphorus 1 10.201 10.201 44.084 < 0.001

Nitrogen:Phosphorus 1 1.245 1.245 5.382 0.0261

Residuals 36 8.33 0.231

Summary N effect P effect

Interaction

NxP

Two-way ANOVA

interpretation

a) Persistent pools (grazed) Yes Yes Yes N and P colimited

b) Ephemeral pools (grazed) Yes Yes Yes N and P colimited

c) Persistent pools (ungrazed) Yes Yes Yes N and P colimited

d) Ephemeral pools (ungrazed) Yes Yes Yes N and P colimited

140

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