errata - au-east.erc.monash.edu.au 1.1 a summary of the main components of the nitrogen cycle ......

Errata

Abstract

P xiv L4 ‘…nitrate to nitrate…’ to ‘…nitrate to nitrite…’

Chapter One

P 12 L3 ‘…due its refractory…’ to ‘…due to its refractory…’

P 26 Periods should be added after the first 2 paragraphs.

Chapter Two

P 40 L3 ‘…a simple a solid-state…’ to ‘…a simple solid-state…’

P 45 L2 ‘…high power requires…’ to ‘…high power requirements…’

P 48 L1 ‘…has a been a…’ to ‘…has been a …’

P 49 L13 ‘…and provide…’ to ‘…and provides…’

P 49 L5 ‘…in-situ measurement total phosphorus…’ to ‘…in-situ total phosphorus …’

P 50 Periods should be added after Paragraphs 2, 3 and 4.

P 53 L17 ‘Samples was collected…’ to ‘Samples were collected …’

P 53 Is Fig. 2.1 correctly depicting the filtration process? This schematic does not accurately

represent the filtration process which is difficult to do in 2 dimensions. There is a description

and photograph available in Figure 2.3.

P 63 L8 ‘…are strongly absorbed…’ to ‘…strongly absorb…’

P 63 L10 ‘…mediums…’ to ‘…media…’

P 68 Fig. 28 ‘Error bar are…’ to ‘Error bars are…’

P 68 L5 ‘…mLmin-1…’ to ‘…mL min-1…’

P 69 It should be explained how the detection limit was determined. The table test describes

the detection limit as being determined by a linear regression method and lists a reference.

P 71 L6 ‘…methods tolerance…’ to ‘…method’s tolerance …’

P 83 Fig. 2.17 ‘…of interested…’ to ‘…of interest…’

P 84 Fig. 2.18 ‘…of interested…’ to ‘…of interest…’

P 84 Is the ‘Waste Treatment Plant’ the Western Treatment Plant of Melbourne Water?

P 84 L6 ‘…waster…’ to ‘…waste…’

P 86 L3‘The data in…indicates…’ to ‘The data in…indicate…’

Chapter Three

P 103 L10 ‘…give rise substantial…’ to ‘…give rise to substantial …’

P 108 L6 ‘The data from…indicates…’ to ‘The data from…indicate…’

P 113 L3/2 ‘…diode diode…’ to ‘…diode…’

P 115 L3 ‘…by dissolved…’ to ‘…by dissolving…’

P 121 L4 ‘…to determine to the extent to…’ to ‘…to determine the extent to …’

P 122 L7 ‘…that that…’ to ‘…that…’

P 124 L3 ‘…that results…’ to ‘…that result…’

P 125 Table 3.4 ‘…and normalized the absorptivity coefficient…’ to ‘…and normalized

absorptivity coefficient…’

P 128 L7 ‘…is relative small…’ to ‘…is relatively small …’

P 130 L3 ‘The data indicates…’ to ‘The data indicate …’

Chapter Four

P 148 L5 ‘…using automated by flow injection analysis…’ to ‘…automated by flow injection

analysis …’

P 50 Periods should be added after the bullet points.

P 168 L2 ‘Figure 4.5 indicates…’ to ‘Figure 4.9 indicates …’

P 169 L12 ‘The data…indicates…’ to ‘The data…indicate…’

P 172 Fig. 4.12 ‘Error bar are…’ to ‘Error bars are…’

P 174 Fig. 4.14 Sensitivity also depends on the signal-to-noise ratio.

P 174 L8 ‘The data…shows…’ to ‘The data…show…’

P 175 Fig. 4.15 ‘…in the presence of 5.0 g L-1…’ to ‘…in the presence of up to 5.0 g L-1…’

P 177 L1 ‘…by a autoclave …’ to ‘…by an autoclave …’

Chapter Five

P 194 L5 ‘…data…was…’ to ‘…data…were …’

P 195 Periods should be added after the first 3 paragraphs.

P 197 Periods should be added after Bullet Points 1 and 3.

P 199 Period should be added after Bullet Point 1.

Copyright Notices Notice 1 Under the Copyright Act 1968, this thesis must be used only under the normal conditions of scholarly fair dealing. In particular no results or conclusions should be extracted from it, nor should it be copied or closely paraphrased in whole or in part without the written consent of the author. Proper written acknowledgement should be made for any assistance obtained from this thesis. Notice 2 I certify that I have made all reasonable efforts to secure copyright permissions for third-party content included in this thesis and have not knowingly added copyright content to my work without the owner's permission.

i

Spectrophotometric Flow Analysis Techniques for the Determination of Total Phosphorus and

Total Nitrogen in Natural Waters

Brady Gentle BSc(2005)

BSc(Hons)(2006)

A thesis submitted in total fulfillment of the requirements for the degree of Doctor of Philosophy

Water Studies Centre School of Chemistry Monash University

Clayton, Victoria, Australia

September 2010

ii

TABLE OF CONTENTS

Table of Figures.......................................................................................................vi

Table of Tables ........................................................................................................xi

List of Publications.................................................................................................xii

Abstract .................................................................................................................xiii

Statement of Authorship.......................................................................................xvi

Acknowledgments ................................................................................................xvii

Abbreviations .........................................................................................................xx

Symbols.................................................................................................................xxii

CHAPTER 1 - INTRODUCTION...........................................................................1

1.1 Introduction........................................................................................................2

1.2 Nitrogen in natural waters .................................................................................4 1.2.1 The aquatic nitrogen cycle .............................................................................4 1.2.2 Nitrogen speciation........................................................................................6

1.3 Phosphorus in natural waters ............................................................................9 1.3.1 The aquatic phosphorus cycle ........................................................................9 1.3.2 Phosphorus speciation .................................................................................11

1.4 Environmental monitoring of nutrients ..........................................................14

1.5 Principles of flow injection analysis.................................................................16 1.5.1 Principles.....................................................................................................16 1.5.2 Dispersive processes....................................................................................18 1.5.3 The refractive index effect ...........................................................................21 1.5.4 Reagent injection flow injection analysis .....................................................23 1.5.5 Portable flow analysis instrumentation.........................................................24

1.6 Research objectives ..........................................................................................25

1.7 References.........................................................................................................27

iii

CHAPTER 2 – A COMPACT PORTABLE FLOW ANALYSIS SYSTEM FOR

THE RAPID DETERMINATION OF TOTAL PHOSPHORUS IN NATURAL

WATERS................................................................................................................37

2.1 Introduction......................................................................................................38 2.1.1 Phosphorus in natural waters .......................................................................38 2.1.2 Techniques for measuring reactive phosphorus in natural waters .................39 2.1.3 Techniques for digestion of total phosphorus in natural waters ....................42 2.1.4 Ozone as an alternate digestion agent to peroxodisulfate..............................47 2.1.5 Flow analysis methods for the in situ determination of phosphorus..............48

2.2 Experimental ....................................................................................................51 2.2.1 Reagents......................................................................................................51 2.2.2 Instrumentation............................................................................................53

2.3 Results and discussion......................................................................................59 2.3.1 Suppression of silicomolybdenum blue interference in total phosphorus measurements.......................................................................................................59 2.3.2 Evaluation of dissolved ozone as a potential oxidant....................................61 2.3.3 Optimisation of digestion conditions for total phosphorus measurement using peroxodisulfate oxidant ........................................................................................63 2.3.4 Analytical figures of merit ...........................................................................68 2.3.5 Laboratory evaluation of the optimised technique ........................................70 2.3.6 Instrumental and method stability ................................................................73 2.3.7 Results of continuous in situ total phosphorus measurement during the Two Bays study............................................................................................................79 2.3.8 Interpretation of the total phosphorus data obtained during the Two Bays cruise ...................................................................................................................82

2.4 Conclusion ........................................................................................................87

2.5 References.........................................................................................................90

CHAPTER 3 – DESIGN AND CONSTRUCTION OF A TOTAL INTERNAL

REFLECTIVE FLOW CELL FOR USE IN FLOW ANALYSIS .......................99

3.1 Introduction....................................................................................................100 3.1.1 Flow cell design.........................................................................................101 3.1.2 Multi-reflective flow cells..........................................................................104 3.1.3 Total internal reflective cells......................................................................106

3.2 Experimental ..................................................................................................113 3.2.1 Design and construction of flow cells ........................................................113 3.2.2 Reagents....................................................................................................115 3.2.3 Flow Injection Apparatus...........................................................................116

iv

3.2.4 Estimated pathlength of the capillary multi-reflective and total internal reflective cell......................................................................................................117

3.3 Results and Discussion ...................................................................................121 3.3.1 Sensitivity of multi-reflective cells and accuracy of the estimated pathlength...........................................................................................................................121 3.3.2 Evaluation of the analytical performance of the z-configuration and reflective cells using the photometric determination of reactive phosphorus.......................126 3.3.3 Comparison of refractive index effects on the total internal reflective, coated multi-reflective and z-cells .................................................................................127

3.4 Conclusion ......................................................................................................131

3.5 References.......................................................................................................133

CHAPTER 4 – ULTRA-VIOLET SPECTROPHOTOMETRIC FLOW

ANALYSIS METHODS FOR THE DETERMINATION OF NITRATE AND

TOTAL NITROGEN IN FRESHWATERS .......................................................136

4.1 Introduction....................................................................................................137 4.1.1 Nitrogen in natural waters..........................................................................137 4.1.2 Techniques for measuring dissolved inorganic nitrogen species in natural waters.................................................................................................................138 4.1.3 Techniques for digestion of total nitrogen..................................................144 4.1.4 Direct measurement of nitrate in the presence of residual peroxodisulfate..149

4.2 Experimental ..................................................................................................154 4.2.1 Reagents....................................................................................................154 4.2.2 Instrumentation..........................................................................................155

4.3 Results and Discussion ...................................................................................159 4.3.1 Interference of chloride for ultra-violet measurement of nitrate..................159 4.3.2 Measurement of nitrate in freshwaters using second derivative spectroscopy...........................................................................................................................160 4.3.3 Interference of residual peroxodisulfate in the ultra-violet measurement of digested total nitrogen ........................................................................................166 4.3.4 Measurement of total nitrogen using second derivative spectroscopy .........171

4.4 Conclusion ......................................................................................................179

4.5 References.......................................................................................................182

CHAPTER 5 – CONCLUSIONS AND FURTHER RESEARCH .....................193

5.1 Introduction....................................................................................................194

5.2 Total phosphorus............................................................................................194

v

5.3 The total internal reflective flow-cell.............................................................196

5.4 Total nitrogen .................................................................................................199

PUBLICATIONS ARISING FROM THE RESEARCH IN THIS THESIS .....201

vi

Table of Figures

Figure 1.1 A summary of the main components of the nitrogen cycle. .......................5

Figure 1.2 Operational classifications of aquatic nitrogen. .........................................7

Figure 1.3 A summary of the main components of the aquatic phosphorus cycle. ....10

Figure 1.4 Operational classifications of aquatic phosphorus. ..................................12

Figure 1.5 The spectrophotometric methods (based on phosphomolybdenum blue

chemistry) used to determine phosphorus speciation.................................................13

Figure 1.6 A typical first generation flow analysis instrument..................................17

Figure 1.7 Schematic representation of dispersion in flow injection analysis ...........19

Figure 1.8 Secondary flow processes within curved tubing......................................20

Figure 1.9 A schematic representing the refractive index effect when using a z-

configuration flow cell .............................................................................................22

Figure 2.1 A schematic representing the total phosphorus analyser. .........................53

Figure 2.2 A labeled picture of the digestion module. ..............................................55

Figure 2.3 A zoomed in and labeled picture of the digestion module. ......................55

Figure 2.4 A schematic of the in-house constructed ozone generator. ......................56

Figure 2.5 A comparison of the analytical response of phosphomolybdenum and

silicomolybdenum blue at 660 nm............................................................................60

Figure 2.6 The relative mineralisation of phytic acid by dissolved ozone and photo-

oxidation with 0, 1 and 2 minute stop times..............................................................62

Figure 2.7 The conversion of sodium tripolyphosphate to reactive phosphorus with

varying sulfuric acid concentration...........................................................................64

Figure 2.8 The change in conversion efficiency with varying peroxodisulfate

concentration............................................................................................................66

vii

Figure 2.9 Oxidation of phytic acid with varying photo-reactor tubing length..........68

Figure 2.10 Replicate peak responses for a blank, 50, 100 and 200 µgPL-1

orthophosphate standards .........................................................................................69

Figure 2.11 A bar chart comparing the total phosphorus concentration as determined

by the flow analysis method and the comparative method ........................................72

Figure 2.12 A comparative line chart indicating the approximate 10 % bias towards

the proposed flow analysis method...........................................................................73

Figure 2.13 The determination of total phosphorus by the flow analysis method over

a two week period ....................................................................................................76

Figure 2.14 The measured phosphorus concentration of a sample over 168 hours as

determined by the proposed flow analysis method....................................................78

Figure 2.15 A map indicating the total phosphorus concentration (5 - 110µgPL-1) as

determined in situ at locations recorded using a GPS unit.........................................80

Figure 2.16 A comparative chart indicating strong agreement with the comparative

method and continuous flow in situ measurements ...................................................81

Figure 2.17 Total phosphorus concentration as determined in situ and plotted against

time for 11-Jan-2010................................................................................................83


time for 23-Jan-2010................................................................................................84


time for 12-Jan-2010................................................................................................85

Figure 3.1 A schematic showing the fundamental design of z-configuration

photometric flow-through cell ................................................................................102

Figure 3.2 A schematic representation of the coated multi-reflective capillary…....104

viii

Figure 3.3 The percentage reflectance of silver and aluminium coating in comparison

to irradiance wavelength and coating thickness. .....................................................105

Figure 3.4 A diagram representing total internal reflection ....................................107

Figure 3.5 An optical simulation of light undergoing total internal reflection within a

circular quartz capillary..........................................................................................110

Figure 3.6 The total internal reflection cell capillary mounted on a metal stand. ....113

Figure 3.7 Flow injection apparatus for phosphorus used to evaluate the performance

of the three cells .....................................................................................................117

Figure 3.8 Flow injection apparatus for the detection of bromothymol blue used to

evaluate the performance of the three cells .............................................................117

Figure 3.9 A representation of light introduction and a single reflection in an

externally coated capillary cell ...............................................................................118

Figure 3.10 A plot of the ratio of estimated optical pathlength to capillary length as a

function of the light beam entry angle taken with respect to the normal of the flow

axis. .......................................................................................................................123

Figure 3.11 The refractive index effect on the z-configuration cell, the circular coated

multi-reflective cell and the circular total internal reflection cell ............................128

Figure 3.12 Flow injection peaks for orthophosphate in nutrient depleted marine

water for the z-configuration cell............................................................................129

Figure 3.13 Flow injections peaks for orthophosphate in nutrient depleted marine for

the total internal reflective cell ...............................................................................130

Figure 4.1 A schematic representing the digestion module and the single reflection

flow-cell detector ...................................................................................................156

Figure 4.2 The single-reflection flow-through cell.................................................157

ix

Figure 4.3 Spectra of various dilutions of artificial sea water and a 1 mgNL-1 as

nitrate solution. ......................................................................................................160

Figure 4.4 The absorbance spectra of a 10.0 mgNL-1 as nitrate standard, with the first

and second derivative also shown...........................................................................161

Figure 4.5 The second derivative spectra of 1 mgNL-1 nitrate and nitrite standards,

along with a standard consisting of a 1:1 mixture of the two...................................162

Figure 4.6 Second derivative peaks for nitrate standards in the 0.0 - 2.0 mgNL-1

range. .....................................................................................................................163

Figure 4.7 A comparison of nitrate concentration as determined by the second

derivative method and a comparative method (cadmium reduction-Griess assay). ..165

Figure 4.8 Ultra-violet spectra of nitrate standards (0 - 2 mgNL-1) with ultrapure

water (UPW) and 2.5 gL-1 alkaline peroxodisulfate (P’sulfate). ..............................166

Figure 4.9 A 2.5 gL-1 peroxodisulfate solution exposed to ultra-violet irradiation from

a medium pressure ultra-violet lamp for 0 - 30 minute periods of time. ..................167

Figure 4.10 Ultra-violet spectra of nitrate standards (0.0 - 2.0 mgNL-1) with 2.5 gL-1

peroxodisulfate after irradiation for 15 minutes. .....................................................168

Figure 4.11 Ultra-violet spectra of nitrate standards (0 – 2 mgNL-1) and a freshwater

sample in 2.5 gL-1 peroxodisulfate are irradiated for 15 minutes.............................170

Figure 4.12 The effect of irradiation time on peroxodisulfate decomposition. ........172

Figure 4.13 Second derivative spectra of residual peroxodisulfate after different

irradiation times. ....................................................................................................173

Figure 4.14 The effect of irradiation time on the sensitivity of second derivative

nitrate detection in the presence of 2.5 gL-1 peroxodisulfate....................................174

x

Figure 4.15 The effect of peroxodisulfate concentration on the sensitivity of direct

ultra-violet and second derivative nitrate detection in the presence of 5.0 gL-1

peroxodisulfate.......................................................................................................175

Figure 4.16 The percentage conversion of various 1 mgNL-1 model compounds ....176

Figure 4.17 A comparison of total nitrogen concentration as determined by the

second derivative method and a comparative method .............................................178

xi

Table of Tables

Table 2.1 The effect of an artificial marine water matrix on the conversion of phytic

acid to orthophosphate .............................................................................................63

Table 2.2 The analytical figures of merit for the total phosphorus flow analysis

system......................................................................................................................69

Table 2.3 Properties of the natural water samples collected, as measured in situ. .....71

Table 3.1 Data indicating the relative transmittance of a silver coated and uncoated

total internally reflective cells ................................................................................108

Table 3.2 Physical properties and optical parameters of the coated capillary multi-

reflective and the total internal reflective cells........................................................120

Table 3.3 Comparison of the sensitivity (calibration gradient) and dispersion of the Z,

capillary multi-reflective and total internal reflective cells......................................122

Table 3.4 Absorptivity values corrected for dispersion and pathlength and normalised

using the absorptivity coefficient of bromothymol blue ..........................................125

Table 3.5 Analytical performance of the three cells for the determination of reactive

phosphorus.............................................................................................................127

Table 4.1 The analytical figures of merit for the second derivative nitrate method. 163

Table 4.2 A summary of the differences in sensitivity and limit of quantification with

increased ultra-violet irradiation time for the detection of nitrate. ...........................169

Table 4.3 The analytical figures of merit for the photo-oxidative total nitrogen

method using second derivative detection...............................................................177

xii

List of Publications

A compact portable flow analysis system for the rapid determination of total

phosphorus in estuarine and marine waters

Brady S. Gentle, Peter S. Ellis, Peter A. Faber, Michael R. Grace, and Ian D.

McKelvie

Analytica Chimica Acta 674 (2010) 117 – 122…………………………………..…202

A versatile total internal reflection photometric detection cell for flow analysis

Peter S. Ellis, Brady S. Gentle, Michael R. Grace, and Ian D. McKelvie

Talanta 79 (2009) 830 – 835………………………………………………………..208

xiii

Abstract

Eutrophication, the over enrichment of nutrients in an aquatic system, is associated

with harmful algal blooms, and as such is a serious environmental issue.

Consequently, interest in the monitoring of nutrient concentrations in aquatic systems

has increased in tandem with a burgeoning public and scientific awareness of

environmental problems. This thesis describes the development of a number of flow

analysis techniques for the monitoring of nutrient concentrations in natural waters;

namely total phosphorus and total nitrogen, as well as the design and construction of a

total internal reflective flow-cell for use in flow analysis systems.

The portable flow analysis system for the determination of total phosphorus in natural

waters was designed with rapid underway monitoring in mind. The digestion module

consisted of a ultra-violet photo-reactor, thermal heating unit, in-line filter and

debubbler, with sample being merged with an acidic peroxodisulfate digestion

reagent. A multi-commutational flow analysis unit was used to introduce gas-

pressurised molybdenum blue chromogenic reagents using two miniaturised solenoid

valves, followed by spectrophotometric detection using a multi-reflective flow cell

with a light emitting diode source and photo-diode detector. The fully automated

system has a throughput of 115 measurements per hour, a detection limit of 1.3

µgPL-1, is highly linear over the calibration range of 0 - 200 µgPL-1 (r2 = 0.9998), and

a precision of 4.6 %RSD at 100 µgPL-1 (n=10). Shipboard field validation of the

instrument and method was performed in Port Philip and Western Port Bays in

Victoria, SE Australia, where 2499 analyses were performed over a 25 hour period,

over a cruise path of 285 kilometres. Good agreement was observed between

xiv

determinations of samples taken manually and analysed in the laboratory and those

measured in situ with the flow analysis system.

Historically, total nitrogen has been determined by Kjeldahl digestion or oxidative

digestion to nitrate followed by reduction of the generated nitrate to nitrate by

cadmium with spectrophotometric detection via the Griess assay. The Kjeldahl

digestion does not measure nitrate and nitrite, and the reduction of nitrate to nitrite

involves the use of a toxic cadmium reagent that degrades rapidly in the presence of

residual oxidant. The flow analysis system developed for the measurement of total

nitrogen involves photo-oxidation of all nitrogenous compounds in the presence of

alkaline peroxodisulfate, with in-line filtration and debubbling, followed by ultra-

violet second derivative spectrophotometric detection of the nitrate generated. A ten

minute stop flow period in the photo-reactor removes a substantial amount of residual

oxidant, which is a spectral interferent in the 220 nm range used to quantify nitrate.

Second derivative spectroscopy is used to minimise interference from any residual

oxidant, as well as other species such as sulfate. The fully automated system has a

throughput of 5 measurements per hour taken in triplicate, has a detection limit of

0.05 mgNL-1, is highly linear over the calibration range of 0 - 2 mgNL-1 (r2 = 0.9989),

and features a precision of 1.2 %RSD for 1 mgNL-1 as ammonia (n = 10). Excellent

agreement was found between storm water samples measured using the flow analysis

system in comparison to those obtained using a reference method.

The design and construction of a total internal reflective photometric flow-through

cell is described. This cell consists of a tubular length of fused silica quartz capillary,

where light is introduced and collected from the cell using quartz optical fibres.

xv

Incident light undergoes total internal reflection at the air-quartz external wall

interface, and thus undergoes multiple reflections as it propagates through the

capillary. This cell was found to have several desirable features in common with

liquid core waveguides (efficient light throughput that leads to high signal to noise

ratio, versatile choice of irradiant light wavelength) and coated capillary multi-

reflective cells (low hydrodynamic dispersion, no entrapment of bubbles, high

tolerance to refractive index effects).

xvi

Statement of Authorship

This thesis contains no material published elsewhere or extracted in whole or in part

from a thesis presented by me for another degree or diploma, except where reference

is made in the text of this thesis.

No other person’s work has been used without due acknowledgment in the main text

of the thesis.

This thesis has not been submitted for the award of any other degree or diploma in

any other tertiary institution.

Brady Gentle

10th September 2010

xvii

Acknowledgments

I would like to acknowledge and express my gratitude to the following people:

Firstly my supervisor, Associate Professor Ian McKelvie. It’s been very beneficial for

me to learn from someone who has an incredible passion for what they do. Ian has

made a concerted effort to invest a large amount of thought and toil into my education

and research, despite often facing heavy time constraints and difficulties of his own,

for which I am very grateful. Ian has always been full of new ideas, useful and

accurate criticism, and valuable advice. Thanks particularly Ian for stressing the

importance of the presentation and communication of scientific results, and patiently

helping me improve my technique in this regard. I would also like to acknowledge Ian

for his financial support during my candidacy.

Many thanks also to Peter Ellis. It’s refreshing to work with someone who is not

afraid of a challenge or problem. It’s been interesting watching and learning from

Peter as he troubleshoots many of the complications that arise daily in the laboratory.

Peter has also been very positive in encouraging my attempts to tackle the more

technical side of my research, and also has a handy knack of demystifying seemingly

very complex things. Thanks also Peter for the enormous time investment into my

research, particularly when that help required you to put in many hours outside of

scheduled work time or when there were more pressing matters that required your

attention.

xviii

Thanks also to Associate Professor Mike Grace for his time and effort, especially in

the months when Ian McKelvie was absent.

The faces that have come and gone in the FIA lab over the years for being helpful and

just general good company: Dr. Barlah Rumhayati, Dr. Asep Saefumillah, Dr.

Elizabeth Reisman, Peter Faber, Yuki O’Bryan and Dr. Ali Shabani.

Thanks to Tina Hines and Kerrie Browne in the WSC analytical lab for their

assistance with the many comparative analyses required in the course of my research.

Thanks as well to Garry McKechnie, Natalie Davey and the crew of SV Pelican 1

during the Two Bays cruise.

My friends who have supported me over the years, either through taking the time to

ask how the “dreaded PhD” is going or just being around for a good laugh: Anwar, the

Bens, Pat, Jeremy, Tim Nam, Dan and Fera. Thanks for taking the time to listen to me

whine and getting involved in my life. I probably couldn’t have dragged myself out of

bed every morning let alone complete this research without you.

A special thanks to Rosemary Sanderson. I can’t imagine how great things would

have been if you were with me from the start.

Thank you to my family, particularly my parents Gail and Steele. Where to start? You

were there from the beginning to the end and for so much more. Thanks for raising me

well and supporting me to reach all my goals.

xix

Thank You to the Lord, my God. This recent time of my life has been the best. You

have created a wonderful and amazing world; thank You for inspiring me to study it

under while living under the roof of Your grace.

xx

Abbreviations

AC Alternating current

A$ Australian dollars

CCD Charge coupled device

DC Direct current

DIN Dissolved inorganic nitrogen

DO Dissolved oxygen

DON Dissolved organic nitrogen

EDTA Ethylenediametetraacetic Acid

FAHP Filterable acid-hydrolysable phosphate

FIA Flow injection analysis

FOP Filterable organic phosphate

FRP Filterable reactive phosphate

HFF Hollow fibre filter

i.d. Internal diameter

LED Light emitting diode

LoD Limit of detection

MRC Multi-reflective cell

NOX The sum concentration of nitrite and nitrate

o.d. Outer diameter

PAHP Particulate acid-hydrolysable phosphate

POP Particulate organic phosphate

PRP Particulate reactive phosphate

rFIA Reverse flow injection analysis

xxi

TAHP Total acid-hydrolysable phosphate

TDN Total dissolved nitrogen

TFP Total filterable phosphate

TIR Total internal reflective cell

TKN Total Kjeldahl nitrogen

TN Total nitrogen

TOP Total organic phosphate

TP Total phosphorus

TPN Total particulate nitrogen

TPP Total particulate phosphate

TRP Total reactive phosphate

UV Ultra-violet

V Volt

xxii

Symbols

A Absorbance

b Optical cell pathlength

c Analyte concentration

C0 Initial concentration of the injected sample zone (prior to dispersion)

Cmax Concentration of dispersed sample zone

D Dispersion coefficient

Dmax Maximum dispersion

k A constant for determining the dispersion coefficient

n Number of injections; refractive index

Po Incident light beam intensity

P Emergent light beam intensity

r2 Coefficient of determination

S Salinity

Sv Injected sample volume

T Transmittance

%RSD Percentage relative standard deviation

σn-1 Standard deviation

ε Molar absorptivity coefficient

θc Critical angle

1

Chapter 1 - Introduction


2

1.1 Introduction

Since the 1960’s there has been increasing concern, from both scientific and public

spheres, regarding water quality issues. Consequently, there has also been emergent

interest in aquatic ecosystem management, as degraded water quality negatively

impacts environmental and human health on aesthetic, functional and economic

levels[1]. Each year a sizeable amount of money is spent in Australia with the intent

of developing informed management strategies based on information gathered from

the chemical analysis of water[2]. In order to gain a broader understanding of aquatic

ecosystems, a variety of factors such as pH, dissolved oxygen concentration, metal

speciation, and nutrient concentration need to be monitored on a regular basis[3].

Eutrophication, meaning “well nourished” in Greek, can be described as an increase

in nutrient concentration within an aquatic system, often followed by proliferation of

photosynthetic activity[4]. One of the features typical of an aquatic system

undergoing eutrophication are algal blooms, which are a well documented cause of

aquatic system degradation[5]. Algal blooms can cause depletion of dissolved oxygen,

both directly through respiration and indirectly via limiting photosynthesis as a result

of sunlight attenuation from the surface bio-film[6]. While eutrophication is a

naturally occurring process, anthropogenic activity can often accelerate these

events[7]. The combined effect of eutrophication is a loss of aquatic biodiversity

through hypoxia and a reduction in the aesthetic value of the afflicted aquatic

system[8].

As an algal bloom is the excessive growth of phytoplankton[9], the development of an

algal bloom requires both sunlight and an excess of nutrients[10]. Nitrogen,


3

phosphorus, carbon and silicon are the most important nutrients for phytoplankton

production[11]. The Redfield ratio (Equation 1.1) provides an indication of the causal

relationship between nutrients, sunlight and algal growth[12]:

106CO2(g) + 16NO3-(aq) + HPO4

2-(aq) + 122H2O(l) + 18H+

(aq) + hν

↓

[C106H263O110N16P](s) + 138O2(g) (1.1)

or simplified as C:N:P = 106:16:1. The forward reaction is photosynthetic;

phytoplankton will raise the dissolved oxygen concentration during daylight hours.

However, the dense biomass of a bloom will significantly reduce dissolved oxygen

content during the night through respiration[13]. Additionally, as phytoplankton cells

die, they sink to the floor of the aquatic system and are consumed by bacteria, a

process which further consumes dissolved oxygen[14]. Furthermore, upon death algae

may release toxins into the water which are harmful to aquatic fauna[15]

If the algal growth of a given system is not limited by sunlight intensity, then

limitation of growth due to insufficient nutrient concentrations may occur[11, 16]. A

“whole-lake” large scale experiment reported by Schindler[17], where controlled

amounts of phosphorus and nitrogen were used to fertilise a lake near Ontario over a

five year period illustrated the importance of these nutrients for primary production,

and also the ratio of phosphorus to nitrogen in controlling both the severity and the

speciation of algal blooms. Nitrogen and phosphorus have been established to be

limiting nutrients on many occasions, and accordingly interest in the concentrations

and behaviours of these nutrients in natural waters is high[11].


4

For these reasons, monitoring of nutrients, such as phosphorus and nitrogen, provides

a cornerstone for the understanding of aquatic systems; in addition to providing

information valuable in the formulation of strategies to deal with, or prevent,

problems such as eutrophication[18, 19].

1.2 Nitrogen in natural waters

Nitrogen is ubiquitous in the environment, comprising 78 % of the atmosphere in the

form of dinitrogen gas, as well as being found terrestrially in several mineralised

forms and various organic compounds[20]. Nitrogen containing species, such as

proteins and nucleic acids, are essential to biological processes. The Redfield equation

describes the ratio of nitrogen to phosphorus for optimal algal growth conditions as

16:1[12].

1.2.1 The aquatic nitrogen cycle

The aquatic nitrogen cycle describes the processes in which nitrogen is converted

between various forms in the aquatic environment. These transformations can be

carried out through both biological and non-biological processes. Nitrogen availability

affects many key ecosystem functions, and as such the nitrogen cycle is of particular

interest. Figure 1.1 summarises these processes.


5

Figure 1.1 A summary of the main components of the nitrogen cycle.

Atmospheric molecular nitrogen readily dissolves in natural waters. While dinitrogen

gas is unreactive due to the high energy requirement to break its triple bond[21],

several species of bacteria and algae, and particularly blue-green algae[6, 7, 21], are

capable of fixing molecular nitrogen. The product of nitrogen fixation via this

mechanism is ammonium[22]. Similarly organic matter, either from anthropogenic

waste or dead biota, is also converted by bacteria into ammonium, a process called

ammonification[11, 16, 23].

In anoxic conditions, ammonium will undergo uptake from bacteria. In the presence

of dissolved oxygen, ammonium can be oxidised to nitrite and then to nitrate, via a


6

bacterially-driven process called nitrification[6]. The oxidation reaction is outlined

below in Equation 1.2:

NH3(aq) + O2(g) � NO2-(aq)

+ 3H+(aq) + 2e-

NO2-(aq) + H2O(l) � NO3

-(aq) + 2H+

(aq) + 2e- (1.2)

Along with ammonification, nitrification is a mineralisation process involving the

complete decomposition of organic nitrogen containing material to bio-available

inorganic nitrogen, which replenishes the nitrogen cycle.

Nitrate can undergo microbial uptake via assimilatory reduction, but only small

amounts of nitrate are consumed in this way[20]. Nitrate produced by nitrification is

typically reduced to dinitrogen gas via denitrification in anoxic conditions.

Denitrifying bacteria use nitrate as an oxidant in the absence of oxygen[6]. The

complete denitrification process can be expressed as a redox reaction, as detailed in

Equation 1.3:

2NO3−

(aq) + 10e− + 12H+(aq) → N2(g) + 6H2O(l) (1.3)

Nitrogen undergoing this process is therefore removed from the aquatic nitrogen cycle

as the molecular nitrogen gas diffuses from surface waters to the atmosphere[24].

1.2.2 Nitrogen speciation

Aqueous nitrogen species can be operationally classified into two broad groups; those

that can pass through a 0.45 µm filter which are classified as dissolved nitrogen, and

those that cannot, which are the particulate fraction[25]. The total nitrogen (TN)


7

content is the sum of both the filterable and non-filterable fractions, and thus

represents the absolute concentration of nitrogen within the water column. Total

dissolved nitrogen (TDN) is typically determined by mineralising the digestible

nitrogen fraction that passes through a membrane filter; whereas total particulate

nitrogen (TPN), due to its increased refractory properties, is often inferred once the

TN and TDN concentrations are known[26]. Figure 1.2 summarises the different

categories of aquatic nitrogen after filtering.

Figure 1.2 Operational classifications of aquatic nitrogen. Dissolved inorganic nitrogen (DIN) includes species such as nitrate, nitrite, and

ammonia. Aside from dissolved molecular dinitrogen, nitrate is the most abundant

nitrogen species found in natural waters[3], and is one of the most bio-available

forms[25]. Apart from eutrophic waters, nitrite is typically found in low

concentrations in natural waters[25], where it usually occurs as an intermediate


8

oxidation state along the path to the reduction of nitrate or the oxidation of ammonia

to nitrate (Equations 1.2-3)[27]. Due to the relative ease of determining nitrate and

nitrite simultaneously, the two species concentrations are often measured concurrently

and reported as NOX (the sum of NO3- and NO2

-)[28]. Ammonia is a bio-available

dissolved inorganic nitrogen species[29]. Aqueous ammonia is basic (pKa = 9.26 at

25oC), and is largely present in natural waters as the ammonium ion[16]. Despite

nitrate, nitrite and ammonia occurring naturally from processes such as dinitrogen gas

fixation, degradation of sediment, decomposition of biotic waste, as well as

atmospheric deposition in the case of ammonia, elevated levels of these species are

most often the effect of anthropogenic activity[30]. Common anthropogenic sources

include effluent from sewage treatment plants, agricultural runoff from the use of

fertilisers[31], and industrial production[32]. Ammonia may also be present as a result

of the decay of organic nitrogen waste species[23] and accordingly, increased

concentrations of ammonium in surface waters may be indicative of domestic

pollution.

Dissolved organic nitrogen (DON) is defined as the difference between TDN and

DIN[33]. This fraction includes amino acids, amines, polypeptides, urea, colloidal

nitrogen and other complex organic compounds. Dissolved organic nitrogen has not

received the attention that inorganic nitrogen has, as many researchers have

considered these species to be biologically inert[34]. However, recent work has

demonstrated that urea and select amino acids are significant nutrient sources for

algae[35], particularly in oligotrophic marine environments where dissolved organic

nitrogen is the dominant form of fixed nitrogen in surface waters[33].


9

1.3 Phosphorus in natural waters

While phosphorus is the eleventh most abundant element in the earth’s crust, it

represents only 0.1 % by mass and is classified as a trace element[36]. Phosphorus is

an important element biologically. It is a key structural component of energy transport

in the Krebs cycle, a major component of cellular membranes in phospholipids, and

assists in the strengthening of bones through calcium phosphate salts[37]. According

to the Redfield ratio, N:P = 16:1, phosphorus is a limiting nutrient with regards to

algal growth[38]

1.3.1 The aquatic phosphorus cycle

In aquatic systems, phosphorus species originate either from natural sources or

anthropogenic point sources and diffuse inputs. Figure 1.3 summarises a simplified

phosphorus cycle within natural waters.


10

Figure 1.3 A summary of the main components of the aquatic phosphorus cycle.

Inorganic phosphorus, whether natural or anthropogenic, is introduced into the system

via external loading. Natural sources include biodegradation of phosphorus-containing

organisms such as algae and dissolution of phosphate minerals[39]. Iron and

aluminum oxyhydroxide containing sediments will adsorb phosphorus once it is

released via weathering effects, which is the major source of non-anthropogenic

bioactive phosphorus[40]. However, with the increasing industrial, agricultural and

domestic uses of fertilizers and detergents, anthropogenic contributions to phosphorus

concentrations in natural waters are dominant. Anthropogenic sources can include

discharge of sewage and industrial effluent, as well as leaching of nutrient rich runoff

from agricultural land[41].


11

Algae and bacteria compete for the uptake of orthophosphate. During periods of

orthophosphate abundance, some algae may store phosphorus in polyphosphate

vesicles[42]; the stored phosphorus is subsequently hydrolysed during periods of

reduced orthophosphate availability[43]. Some labile dissolved organic species may

also be utilised by bacteria, but many organic compounds possess refractory

properties[44]. Upon the decay of dead algae, or excretion by the larger animals that

consume them, organic phosphorus is released into the water column. Organic

phosphorus may undergo sedimentation, where it can either reenter the system as

inorganic phosphorus via dissolution and decomposition, or bind to the sediment.

1.3.2 Phosphorus speciation

Aquatic phosphorus can be functionally categorised into two macro groups; namely

species that can pass through a membrane filter which are called filterable phosphorus

and those that cannot, i.e. particulate phosphorus[45]. Total phosphorus (TP) is the

measurement of the sum of all phosphorus containing compounds and includes

phosphorus that exists in colloidal and particulate matter, within organisms and

dissolved in waters[46]. Figure 1.4 shows the fractions and operational distinctions of

phosphorus once separated by a membrane filter.

Total filterable phosphorus (TFP) can be further classified into several subgroups.

Filterable inorganic phosphorus includes orthophosphate and condensed phosphates.

Orthophosphate is produced by natural processes and is found in natural waters,

sediments and effluent. Filterable reactive phosphorus (FRP) is a practical definition,

and is the filterable fraction of phosphorus which will react with acidic molybdate to


12

form phosphomolybdate[47]. It is important to note that this fraction may contain a

measure of labile dissolved colloidal, organic and condensed phosphates[48], and thus

does not represent orthophosphate exclusively, which is the most bio-available

fraction of phosphorus[49]. Condensed phosphates, which are generally found in

waste waters in high concentrations as polyphosphates, metaphosphates, and branched

ring structures[46], degrade into orthophosphate over the course of hours[41], and

consequently occur only in low concentration in natural waters. Filterable condensed

and colloidal phosphates are often collectively referred to as acid-hydrolysable

phosphate (FAHP), as they readily undergo hydrolysis to orthophosphate in the

presence of acid.

Figure 1.4 Operational classifications of aquatic phosphorus[45].

Filterable organic phosphorus (FOP) includes various biologically produced species

such as nucleic acids, phospholipids and phosphoproteins, as well as colloidal


13

phosphorus and organic condensed species[50]. Common organic phosphorus species

include adenosine triphosphate, phospholipids and inositol phosphates[45].

The particulate phosphorus fraction includes inorganic and organic phosphates bound

to clays and heavy colloidal matter, as well as phosphorus contained within biota.

Because of the difficulty of determining particulate phosphorus due its refractory

properties, this value is often inferred from the total dissolved phosphorus and total

phosphorus measurements[51]. Figure 1.5 schematically represents the

spectrophotometric methods for determining the concentration of various aquatic

phosphorus fractions.

Figure 1.5 The spectrophotometric methods based on phosphomolybdenum blue chemistry used to determine phosphorus speciation. The particulate fraction can be determined by subtracting the filterable fraction from the total fraction above it, as represented in Figure 1.4. TOP is determined via subtraction of the other three known values in the total phosphorus pool.

TRP TAHP + TRP TOP

FRP FAHP + FRP FOP + FRP

TP

TFP

Sample

Molyb. Blue

Molyb. Blue

Conc. H2SO4 Molyb. Blue

Conc. H2SO4 Molyb. Blue

Alk. digestion Molyb. Blue

Digestion + Acid Molyb. Blue

TP – (TRP + TAHP) =

Digestion + Acid Molyb. Blue

Filtration


14

As indicated in Figures 1.4 and 1.5, total phosphorus can be considered a

measurement of the maximum potential bioavailable phosphorus, whereas filterable

reactive phosphorus, consisting mainly of orthophosphate, gives an indication of the

most readily bioavailable phosphorus.

1.4 Environmental monitoring of nutrients

Monitoring environmental parameters is a cornerstone to understanding aquatic

ecosystems, and a key step in developing management strategies[3]. Thus, there has

been an increasing interest in the development of reliable techniques for quantification

of environmental factors, such as nutrient concentration.

The monitoring of nutrients in natural waters has been conventionally achieved via

periodic manual sampling, freezing and transport of the collected sample, followed by

laboratory-based analysis[52]. From an ease of operation standpoint, laboratory-based

analysis is simpler compared with field analysis, in addition to offering greater control

over external factors such as temperature, stability of power source, vibration, and so

on. Nevertheless, manual sample collection, transportation and storage present the

following difficulties;

• the chemical matrix of the sample may degrade with time, due to either

chemical or microbial reaction, improper storage, or contamination [26, 53,

54]

• microbial activity may alter the sample during sample filtration or through

contamination existing in the storage vessel[55]


15

• manual sampling, transportation, and subsequent laboratory analysis are

expensive in terms of both time and cost, which discourages intensive spatial

or temporal sampling[56]

• due to the dynamic nature of aquatic systems, the delay between sampling and

laboratory analysis may cause the data to be unrepresentative of the sampled

area’s current condition[57].

Measurement of sample in situ eliminates the need for sample transportation and

storage, and thus reduces the chance of contamination or degradation, and also

reduces the time lag between the sample measurement and collection, resulting in

improved data quality and relevance[52]. Field instrumentation constructed to collect

and treat sample automatically also reduces the risk of operator error or

contamination.

There are many commercially available fixed-site monitoring systems based on ion

selective electrodes[58], fibre optic probes and biosensors[59] for the quantification of

various chemical species in natural waters. However, many of these sensors suffer

from poor selectivity and sensitivity, and thus their applications in natural waters will

be limited[11, 53, 57]. Automated chemical analysers, which include flow injection

analysis techniques, differ from sensors in that sample is reacted with reagents in a

controlled manner, followed by interrogation via an appropriate detector. These

systems offer more long term stability than sensors, and have the additional capability

of performing in situ re-calibrations[54].

Flow injection analysis (FIA), a versatile technique for the handling of liquid phase

chemical analyses, is well suited for in-field water monitoring[60]. Flow analysis


16

techniques offer the potential for high throughput (tens to hundreds of measurements

per hour) enabling the provision of data of high temporal and spatial resolution. This

rapid response allows the resolution of chemical “hot spots” such as effluent

discharge points or other areas of point source nutrient input, as well as the possibility

of highly detailed surface water mapping[61]. Other features of flow analysis systems

desirable for field operation include robustness, simple operation and maintenance,

portability and low reagent consumption in the case of reverse flow injection.

1.5 Principles of flow injection analysis

The first reported FIA technique was by Ruzicka and Hansen in 1975[62]. Over the

succeeding 35 years, there has been an ever increasing amount of research and

analysis undertaken using flow injection techniques[63]. Since its inception flow

injection techniques have become routine in laboratories[53, 64], and have proven

particularly useful in the field of nutrient analysis in waters[65].

1.5.1 Principles

The first generation of flow analysis techniques, in their simplest form, involved the

injection of a small volume of sample into a continuously flowing unsegmented

carrier stream, which could be later merged with suitable reagents[60]. The flow was

typically regulated by a peristaltic pump, with the solution flowing through narrow

bore tubing (generally 0.3 to 1.0 mm i.d.). The sample is reproducibly injected into

the flowing carrier stream using a valve with a corresponding sample loop that is


17

refilled subsequent to each measurement. As the sample zone flows downstream, the

solution undergoes dispersive processes that cause the carrier solution and the injected

solution to mix. Confluence points allow for the addition of reagent streams. As the

mixture further disperses following the addition of reagents, product will begin to

form at the sample and reagent interface. The mixing process can be further enhanced

by the addition of mixing devices, which may be simple knotted or coiled lengths of

tubing, or more complicated devices such as mixing chambers. A detector placed at

the end of the stream will then measure a parameter that changes upon the injection of

the sample zone. The detector response takes the form of a peak, the height and area

of which is directly proportional to the analyte concentration of the injected liquid

sample. Figure 1.6 is a representation of a typical first generation FIA instrument.

Figure 1.6 A typical first generation flow analysis instrument[60].

The information recorded by the detector is thus a result of two processes occurring in

the flowing sample stream; namely the rate of the ensuing chemical reactions and the

sample zone dispersion. For successful operation, all samples measured by the flow

injection instrument must be processed in exactly the same manner. The sample

injection volume, the flow rate, and any measures taken to enhance mixing, must be

highly reproducible. It is important to note that because of the high level of control


18

over these conditions offered by flow injection techniques, the chemical reactions

involved need not reach equilibrium in order to provide precise analytical responses.

Propulsion using a peristaltic pump can cause the flow to oscillate due to the nature of

the pumping, which can cause fluctuations at the detector. However, this effect is

severely reduced as the amount of longitudinal dispersion increases within the

manifold. This means that the longer the length of tubing used, the more the pulsing is

dampened. In most flow injection manifolds, the length of tubing used will be

sufficient enough that the effect of this pulsing will be so small as to not cause a

significant reduction of reproducibility[66]. Alternative methods of propulsion to

peristaltic pumps include syringes, piston pumps, gravimetric flow, and pressurised

pumping devices[60, 64, 67].

1.5.2 Dispersive processes

Sample injection is commonly performed using a rotary injection valve, where sample

is loaded into a fixed-volume loop of tubing prior to injection[68]. Alternatively,

sample can be introduced to the carrier stream using time-based switching, which may

be achieved using solenoid valves or a syringe pump[69]. After injection, the sample

zone flows under laminar flow conditions[70], and will mix with the carrier stream, or

reagent streams if in use, via axial and radial dispersive processes (Figure 1.7). If the

sample injection volume is too large, an un-mixed zone in the centre of the sample

plug will occur, which causes a double peak to be recorded.


19

Figure 1.7 Schematic representation of dispersion in flow injection analysis. (a) is radial dispersion and (b) is axial dispersion. The sample zone will disperse axially, causing the injection zone to broaden along the

length of the tubing. Axial dispersion is undesirable as it can cause peak broadening,

which will result in a loss of sensitivity and a reduction of sample throughput[71].

Due to the parabolic laminar velocity profile of the front and rear interfaces of the

sample zone radial dispersion will occur, which assists in the formation of product at

the interface. Radial dispersion is a combination of molecular diffusion and secondary

flow, which is caused by centripetal forces on the fluid as it flows around a bend with

the strength of the force being dependent upon fluid velocity and curve radius[72].

The leading interface of the sample zone continuously diffuses laterally into more-

slowly-moving liquid and the trailing interface diffuses medially into faster-moving

liquid (Figure 1.8). Mixing coils are used to enhance secondary mixing along with

radial dispersion. Coil geometries and internal diameters both affect the amount of

dispersion and secondary flow mixing. Figure 1.8 shows schematically the secondary

mixing process.


20

Figure 1.8 Secondary flow processes within curved tubing. Centripetal forces experienced by the fluid as it flows through a bend cause counter rotary flow in the cross sectional plane, leading to sample and reagent mixing. Reproduced from Ruzicka and Hansen[60].

The extent of dispersion is defined by the dispersion coefficient, D, in Equation

1.4[60]:

D = C0 / Cmax (1.4)

where C0 is the concentration of the sample prior to mixing-dispersion and Cmax is the

concentration subsequent to mixing-dispersion. D is directly related to the dilution

that the sample zone experiences. The dispersion coefficient of a given flow analysis

system is often determined practically by measuring the absorbance of a defined

volume of a chromophore before (C0) and after it passes through the instrument

(Cmax)[60]. The maximum dispersion experienced by the sample zone is related to the

size of the zone by Equation 1.5[61]:

1/Dmax = 1 – e-kSv (1.5)

Where Dmax = maximum dispersion, k = constant, Sv = the volume of the injected

sample. The above equation is only applicable to a single line system.


21

Dispersion is a controlling parameter with regard to the sample throughput and

analytical sensitivity of a given FIA system[73]. If the original composition of the

sample solution is to be measured (i.e pH, conductivity etc.), actively limiting

dispersion will yield a maximum analytical response. If however one or more

chemical reactions are required, then increasing dispersion (particularly radial

dispersion), and hence mixing between the sample zone and reagents, will increase

sensitivity up to a point. Johnson et al [74] reported that for the conventional flow

injection analysis of phosphorus, involving the addition of two colorimetric reagents,

a dispersion coefficient of 3 produced the optimum analyte peak.

1.5.3 The refractive index effect

When a sample of high ionic strength is injected into a carrier stream of low ionic

strength, a parabolic lens will occur at the laminar flow profile interface between the

two zones due to the different refractive indices of the two liquids. When photometric

detection is being employed, this parabolic lens causes aberrations in the light path

traversing the flow-through cell. This is called the refractive index, or Schlieren,

effect.

If the refractive index of the injected sample is greater than that of the carrier, the first

parabolic interface to pass through the detector will focus light at the detector surface,

which causes the measured signal to momentarily increase (P) compared with the

incident light intensity (Po), resulting in a reduction in absorbance (A) as according to

Equation 1.6[75]:


22

A = log Po/P (1.6)

As the second parabolic interface passes through the detector, light is dispersed from

the detector surface causing a decrease in the measured signal, which results in an

increased absorbance. Figure 1.9 illustrates the refractive index effect in a z-

configuration flow through cell, for the situation where the refractive index of the

sample is greater than the carrier stream.

Figure 1.9 A schematic representing the refractive index effect when using a z-configuration flow cell. The upper diagram shows the first parabolic convex interface of the sample zone with the carrier zone and the subsequent light focus. The middle diagram shows the second parabolic concave interface of the sample zone with the carrier zone and the subsequent light dispersion. The lower diagram is the light scattering as caused by imperfect mixing i.e. a heterogeneous zone. Reproduced from [76].

P0 P

P

P P0

P0


23

Under the conditions described, the refractive index effect causes a negative peak,

followed by a positive peak, as shown in Figure 1.9. This effect is superimposed upon

the analyte peak, and can alter the recorded peak shape from that expected from the

analyte alone. The magnitude of this effect is exacerbated by increasing the difference

in refractive index across the interface, the degree of sample zone dispersion and the

photometric flow cell design[77]. The refractive index effect may also adversely

affect the detection limit of photometric flow injection methods, and may lead to

substantial errors if left uncorrected[78].

The refractive index effect can either be reduced, or corrected for, using several

methods. Matrix matching the ionic strength of the carrier solution to the sample[79],

correction for the light aberration using dual wavelength detection[78], or introduction

of the light beam transverse to the flow axis (thus significantly reducing the effect of

the parabolic lens[80]) have all been applied. The matrix matching technique is only

effective if the ionic strength of the sample is known and remains constant between

measurements. Therefore, the refractive index effect has proved a significant

hindrance towards the application of flow injection analysis methods in waters with

varying ionic strengths e.g. estuaries[81].

1.5.4 Reagent injection flow injection analysis

Conventional flow injection analysis involves the injection of a small volume of

sample into constantly flowing carrier and reagent streams[60]. While this

configuration may be advantageous for laboratory based measurements where small

amounts of sample are analysed, this arrangement also requires copious amounts of


24

possibly expensive reagents, and produces a correspondingly large amount of waste,

which may be noxious or costly to dispose.

Reverse flow injection analysis (rFIA), sometimes referred to as reagent injection

analysis or multicommutation, involves the injection of small volumes of reagent into

a continuously flowing sample stream[61]. With the elimination of the requirement of

a carrier stream and the adoption of a reagent injection protocol, reverse flow

injection analysis dramatically reduces reagent use[82], as well as achieving a

resultant reduction in waste. For rFIA methods, the amount of sample present in the

reactive zone will increase as dispersion increases[83], and thus reagent injection

arguably has a higher theoretical sensitivity than conventional flow injection analysis

techniques[61].

Reverse FIA is particularly suited to environmental in situ monitoring where large

amounts of sample are available on-site[53], and economical use of reagents is highly

desirable. The multi-channel peristaltic pumps used in conventional FIA are often

bulky and use large amounts of power, whereas miniature components that are highly

compatible with reagent injection protocols are far more economical, both space and

resource wise, than conventional instrumentation[84].

1.5.5 Portable flow analysis instrumentation

With the recent development of miniaturised components for FIA, such as solenoid

valves, syringe pumps, solid-state light emitting diodes and miniaturised charge

coupled devices[85], coupled with advances in computing and associated software,


25

portable flow injection analysis instrumentation have become increasingly feasible

and attractive for in situ water monitoring. Portable flow injection analysis

instrumentation for the in situ analysis of nitrate[52, 86] and underway analysis of

filterable reactive phosphorus[84] have been operated successfully in field conditions.

The development of specialised flow injection analysis instrumentation for underway

analysis of waterways is a step towards the collection of large amounts of high quality

data. These instruments, along with other in-line sensors, are critical in procuring

essential data for developing, and improving existing, management strategies for

natural waters.

1.6 Research objectives

The endeavour of the work described in this thesis is to develop sensitive, accurate,

precise and rapid flow analysis techniques for the determination of total phosphorus

and total nitrogen for both laboratory and field applications. This research will

primarily focus upon the determination of these species in natural waters (fresh,

estuarine and marine), and as such will need to be able to determine part per billion

(µgL-1) concentrations, and handle a variety of complex sample matrices. The major

objectives are:

• To develop a portable flow analysis method for the in situ determination of

total phosphorus, using a peroxodisulfate oxidising medium coupled with a

photo-oxidation and acidic thermal digestion. Photometric detection will be


26

achieved via the molybdenum blue method. The substitution of dissolved

ozone for peroxodisulfate as an oxidant is also investigated

• To develop a flow analysis method for the determination of total nitrogen,

using in-line photo-oxidation of nirogenous species to nitrate. Direct ultra-

violet detection of nitrate is investigated as an alternative to the commonly

used, but toxic, cadmium reduction column

• To design, evaluate and apply a novel total internal reflective flow-through

photometric cell suitable for spectrophotometric flow injection applications. A

comparison of certain characteristics (sensitivity, dispersion, tendency to trap

bubbles, and susceptibility to refractive index effects) is made with a

commercially available z-configuration cell and a coated multi-reflective cell.


27

1.7 References

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contemporary practical problem. American Scientist 61, 269-279.

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3. Robards, K., McKelvie, I.D., Benson, R.L., Worsfold, P.J., Blundell, N.J., and

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8. Ikeda, S., and Adachi, N. (1976). Dynamics of the nitrogen cycle in a lake and

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28

10. Hautier, Y., Niklaus, P.A., and Hector, A. (2009). Competition for light causes

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29

21. Stumm, W., and Morgan, J.J. (1996). Pollution by nitrogen compounds. In

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biological nitrogen removal of extremely ammonia concentrated wastewater.

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25. Oms, M.T., Cerda, A., and Cerda, V. (2000). Analysis of nitrates and nitrites.

In Handbook of water analysis, L.M. Nollet, ed. (New York: Marcel Dekker

Inc.), pp. 201-221.

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and I.D. McKelvie, eds. (Elsevier), p. 335.

77. Bezerra dos Santos, S.R., Ugulino de Araujo, M.C., and Barbosa, R.A. (2002).

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35

schlieren effect measurements using an LED-photocolorimeter. Analyst 127,

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78. Liu, H., and Dasgupta, P.K. (1994). Dual-wavelength photometry with light

emitting diodes. Compensation of refractive index and turbidity effects in flow

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Elimination of the schlieren effect in the determination of reactive phosphorus

in estuarine waters by flow-injection analysis Analytica Chimica Acta 351,

265-271.

80. Ellis, P.E., Lyddy-Meaney, A.J., Worsfold, P.J., and McKelvie, I.D. (2003).

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36

85. Hanrahan, G., Gledhill, M., Fletcher, P.J., and Worsfold, P.J. (2001). High

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Automated Methods & Management in Chemistry 22, 97-102.

37

Chapter 2 – A compact portable flow analysis

system for the rapid determination of total

phosphorus in natural waters

Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters

38

2.1 Introduction

2.1.1 Phosphorus in natural waters

Phosphorus is an essential nutrient for aquatic photosynthetic organisms, particularly

algae[1]. Increased concentrations of phosphorus can lead to eutrophication, and

associated incidences of harmful algal blooms[2]. Phosphorus may be growth limiting

in natural waters, particularly in freshwaters[3]. Heightened occurrences of algal

blooms have been linked to increased anthropogenic phosphorus input, from sources

such as agricultural and domestic run-off, as well as industrial and domestic sewage

effluents[4]. Phosphorus is commonly monitored in aquatic systems[5], both to

examine naturally occurring nutrient cycling and to determine the effects of

anthropogenic activities.

Two commonly measured operational categories of phosphorus are filterable reactive

phosphorus (FRP) and total phosphorus (TP)[5]. Filterable reactive phosphorus is the

fraction that will pass through a 0.45 µm filter and will also react readily with acidic

molybdate to form a heteropoly acid[6, 7]. This category consists primarily of

orthophosphate, which is considered to be the most bioavailable form of aquatic

phosphorus[8], and other filterable, acid hydrolysable species[9]. Total phosphorus is

the measurement of all phosphorus within a body of water; phosphorus within

colloidal and particulate matter, within organisms and dissolved in waters[5].

Measurement of filterable reactive phosphorus gives an estimate of the amount of

bioavailable phosphorus within an aquatic system; determination of the total


39

phosphorus concentration is a measurement of the maximum potential mass of

bioavailable phosphorus.

Total phosphorus concentrations can often be less than 10 µgPL-1 in coastal and open

marine waters and pristine freshwaters, and in excess of 100 µgPL-1 in systems

heavily affected by anthropogenic contamination[10]. ANZECC guidelines

recommend 10 - 100 µgPL-1 as total phosphorus for rivers and streams, and 1 - 10

µgPL-1 as phosphate-phosphorus for marine and coastal waters[11].

2.1.2 Techniques for measuring reactive phosphorus in natural waters

Numerous techniques have been investigated for measuring reactive phosphorus

concentrations in waters, including potentiometry[12], fluorescence spectrometry[13],

atomic absorption spectrometry[14], X-ray fluorescence[15], neutron activation

analysis[16], bioassay [17], as well as colorimetric methods employing the

chromogenic molybdenum blue reaction[18], and also the ion-pair of

phosphomolybdate and malachite green[19]. Potentiometric methods are reported to

have a detection limit of 30 µgPL-1[12, 20] that is too insensitive for use in marine or

pristine waters, and ions commonly found in natural waters such as chloride,

metasilicate and calcium are significant interferents[12]. Bioassay methods are time

consuming, labour intensive, and insensitive. X-ray, atomic absorption and neutron

activation techniques both entail instrumentation that is too large for convenient field

use and generally lack sensitivity.


40

For both field and laboratory based FRP analysis, the molybdenum blue method is the

favoured approach because of its selectivity, sensitivity, and simplicity[2]. In aqueous

conditions, molybdate will form a yellow heteropoly acid complex with

orthophosphate[21], as expressed in Equation 2.1;

HPO42-

(aq) + 12MoVIO42-

(aq) + 26H+(aq) � H3(PMoVI

12O40)(aq) + 12H2O(l) (2.1)

It should be noted that this reaction is pH dependent and is favoured by acidic

conditions. The heteropoly acid is also labile[22], Equation 2.2;

H3(PMoVI12O40)(aq)� 12MoVIO3(aq) + H2PO4

-(aq) + H+

(aq) (2.2)

While the phosphomolybdenum yellow complex can be detected photometrically, the

heteropoly acid may be reduced to form phosphomolybdenum blue, as indicated in

Equation 2.3;

H3(PMoVI12O40)(aq) + 4e- � H3(PMoV

4MoVI8O40

)4−(aq) (2.3)

This reduction is typically achieved by either tin(II) chloride, ascorbic acid, or another

suitable reductant and results in a product with a much greater absorptivity than

phosphomolybdate. The phosphomolybdenum blue complex generated via reduction

by tin(II) chloride has an absorbance maximum of 690 - 700 nm[23] and is sometimes

used in preference to ascorbic acid reduction because it offers a faster rate of

reduction[2]. Furthermore, the wavelength absorbance maxima of 690 - 700 nm

(compared with 880 nm for ascorbic acid[2]) allows the use of a simple a solid-state

detector that utilises a red light-emitting diode as a light source[24]. However, high

concentrations of chloride anions are known to suppress the sensitivity of the tin(II)-


41

reduced phosphomolybdenum blue complex by as much as 13 % in saline marine

waters[7].

Water samples may contain some easily hydrolysable organic and condensed

phosphates that can mineralise to orthophosphate under acidic conditions.

Consequently, the molybdenum blue method is known to overestimate the

concentration of bio-available orthophosphate because of hydrolysis that occurs under

the acidic reaction conditions (Equation 2.1)[25]. While this overestimation poses a

problem when attempting to measure bioavailable phosphorus only, it is beneficial

when measuring total phosphorus as all colloidal and condensed phosphates must be

hydrolysed prior to detection.

Silicon, arsenic and germanium (as orthosilicate, arsenate and germanate) may also

form heteropoly acid complexes with molybdate, and can be present in natural waters

at high enough concentrations to cause significant interference in phosphorus

determination[26]. However, if the reaction conditions are maintained at below pH

one, the kinetics of heteropoly acid complex formation with these interfering species

can be markedly suppressed to the extent that they are almost negligible with

comparison to the formation of phosphomolybdate[26]. A reaction pH of less than one

also ensures that auto-reduction of MoVI to MoV does not occur, which would

otherwise cause the formation of a heteropoly blue product leading to an elevated

blank signal[2].


42

2.1.3 Techniques for digestion of total phosphorus in natural waters

Particulate and some filterable organic and condensed phosphorus species cannot be

measured directly as phosphomolybdenum blue because of their refractory nature, and

hence all phosphorus-containing compounds must first be converted to the more

reactive orthophosphate before conversion to phosphomolybdenum blue. This process

is termed digestion, which may involve dissolution, oxidation or hydrolysis depending

on the nature of the sample. For flow analysis, digestion is usually achieved in-line by

oxidation or hydrolysis, or a combination of both. Subsequent to complete digestion

of all phosphorus containing compounds to orthophosphate, spectrophotometric

measurement of reactive phosphorus is used for quantification of the total phosphorus

concentration.

Natural waters, sediments and sediment pore waters contain phosphorus species of

varying lability. The more refractory organic and particulate phosphorus compounds

may require strong hydrolysing and oxidative conditions in order to undergo complete

mineralisation. The digestion of these compounds may be performed by; thermal

methods (wet chemical digestion[27, 28], high temperature combustion[29, 30],

microwave digestion[31-33]), photo-oxidative methods[34-36], and combined thermal

hydrolytic and photo-oxidative methods[10, 37]. The use of an ultra-violet photo-

oxidative digestion procedure alone is insufficient to convert condensed phosphate

species into orthophosphate, and as such provides the basis for discrimination

between organic and condensed phosphorus fractions[36]. An enzymatic method for

the determination of total phosphorus in cereals has also been described using

enzymatic hydrolysis via the enzyme phytase[38]. However, this approach is


43

unsuitable for hydrolysis of the wide variety of phosphorus species found in natural

and waste waters[39].

An increasingly used method for total phosphorus digestion involves the use of a

peroxodisulfate oxidising medium [10, 31, 32, 35, 37, 40-44]. McKelvie et al[35]

found that rapid photo-oxidation of dissolved organic phosphorus could be achieved

using an alkaline 40 gL-1 peroxodisulfate solution.

Peroxodisulfate, while a strong oxidant, reacts slowly with many organic species[45],

but upon exposure to ultra-violet radiation, a peroxodisulfate medium will produce

hydroxyl and sulfate radicals, which are very strong oxidising agents. Hydrogen

peroxide will also generate hydroxyl radicals upon exposure to ultra-violet

radiation[34, 46]. However, while hydrogen peroxide may find application in

segmented flow methods, it is rarely used in flow injection applications due to the

development of copious amounts of oxygen bubbles upon exposure to ultra-violet

light, which severely impedes spectrophotometric detection.

Hydroxyl and sulfate radical production from peroxodisulfate can be enhanced using

ultra-violet light [35, 37, 40-42, 44], as shown by Equation 2.4[45]:

S2O82-

(aq) + hν → 2SO4-•

(aq)

SO4-•

(aq) + H2O(l) � HSO4-(aq) + OH•

(aq) (2.4)

The hydroxyl and sulfate radicals may then react with organic compounds, further

decompose peroxodisulfate[47](Equation 2.5), or undergo further radical

reactions[6](Equation 2.6):


44

S2O82-

(aq) + OH•(aq) � HSO4

-(aq) + SO4

-•(aq) + 1/2O2(g) (2.5)

SO4-•


-(aq) + 1/2O2(g) (2.6)

Both the sulfate and hydroxyl radicals are responsible for the destruction of organic

compounds. Either radical may dominate this process dependant upon pH, with the

hydroxyl radical being produced principally under alkaline conditions and sulfate

radical production occurring primarily under acidic conditions[45]. Because of this, a

peroxodisulfate ultra-violet method can operate successfully under either acidic[37] or

alkaline[10] conditions.

An alkaline peroxodisulfate medium will selectively digest organic phosphorus

compounds[36]. One advantage of using an alkaline medium is that carbon dioxide

generated by the oxidation of organic matter is suppressed, with the oxidised carbon

present predominantly in the carbonate form, therefore preventing the formation of

carbon dioxide gas bubbles; although oxygen formation from the photo-degradation

of peroxodisulfate will still be prevalent given the relative concentrations of organic

matter and peroxodisulfate. Peat et al[48] found that mineralisation of dissolved

organic phosphorus compounds was significantly suppressed in marine waters under

alkaline digestion conditions. Aminot and Kerouel[46] found that superior recovery of

refractory phosphorus compounds was achieved when the sample salinity was 15 or

less when using a neutral hydrogen peroxide oxidising agent. Magnesium and calcium

ions, which occur commonly in natural waters, may form stable complexes with

organic phosphorus compounds under alkaline conditions[48], which results in

reduced conversion efficiency. Rumhayati[49] found that near complete conversion of


45

phytic acid using photo-oxidation could be achieved in marine waters when the

aforementioned ions were removed using ion-exchange.

In order to mineralise condensed phosphate species, assist in the breakdown of

colloidal or particulate-bound phosphorus, and to avoid precipitation of magnesium or

calcium phosphates[48], acidification of the sample is required. This may be achieved

by either using an acidic peroxodisulfate oxidising medium[32, 37, 41], by allowing

the peroxodisulfate to thermally decompose to generate acid[31], or by a use of an

alkaline medium with an additional in-line acidification step[10].

Aoyagi et al[40] developed an FI system using a 10 m long coiled Teflon capillary

digester, featuring a length of platinum wire inserted into the tubing. The platinum

wire acted as a catalyst for the mineralisation of phosphorus compounds using

peroxodisulfate. The method utilised colour formation of the ion pair of

phosphomolybdate and malachite green to give a low detection limit of 2 µgPL-1;

however, the lengthy capillary reactor limited the sample throughput to one analysis

per four minutes. Hinkamp and Schwedt[32] used a 5 - 10 m crocheted Teflon tubing

reactor in a conventional microwave oven to assist digestion with acidic

peroxodisulfate. A sample throughput of 20 per hour was achieved. The mineralized

phosphorus was detected using an amperometric method, with a modest detection

limit of 100 µgPL-1. Hinkamp and Schwedt[32] also found that lower pH conditions

caused copious bubble formation. Thermal digestion methods involving microwave

heating will find limited field application due to the high power requires of such

systems, e.g. a 650 W microwave oven was used by Hinkamp and Schwedt[32].


46

Pérez-Ruiz et al[10] found that when using a 40 gL-1 peroxodisulfate reagent photo-

oxidation from a low pressure mercury lamp, total mineralisation of organic

phosphorus compounds could be achieved in as little as 20 seconds. However, 60

minutes of heated digestion using 0.5 M hydrochloric acid was required to achieve

similar results with tripolyphosphate and trimetaphosphate model condensed

phosphorus compounds. Using similar methods, Benson et al[37] and Higuchi et

al[42] showed that long digestion times were required for complete mineralisation of

condensed phosphates using high concentrations of acid (0.3 M perchloric acid and

0.2 M sulfuric acid respectively). Reducing the pH of the sample below 0.5 pH using

high concentrations of acid in conjunction with molybdenum blue detection may

suppress phosphomolybdenum blue formation[50], which decreases the sensitivity

and consequently the limit of detection. Benson et al[37] reported a detection limit of

0.15 mgPL-1 and Fernandes and Lima[41] achieved 1 mgPL-1 when using 2 M sulfuric

acid.

Digestion efficiency can be determined by quantifying the percentage mineralisation

of refractory compounds. Many reports indicate that the peroxodisulfate medium

when coupled with photo oxidative treatment can produce conversion efficiencies

equal, or close to, 100 % for refractory organic phosphorus compounds such as phytic

acid[37, 40, 43], especially if a period of stop flow is incorporated into the flow

injection procedure[10]. However, often these methods will exhibit a lower

conversion efficiency for condensed phosphate species[10, 37, 42]. But, this is often

ignored because the proportion of phosphorus present as condensed phosphates is

relatively small in comparison to organic phosphorus and reactive phosphorus species

in natural and waste waters[42].


47

2.1.4 Ozone as an alternate digestion agent to peroxodisulfate

An alternative means of generating hydroxyl radicals is the irradiation of dissolved

ozone. The use of ozone as an oxidant in the digestion of phosphorus compounds for

total phosphorus determination has not previously been reported. Dissolved ozone

degrades in the presence of UV radiation according to Equation 2.7[51]:

O3(aq) + H2O(l) + hν → H2O2(aq) + O2(g)

H2O2(aq) + hν → 2OH•(aq) (2.7)

The hydroxyl radical generated will readily oxidize any organic phosphorus

containing compounds[52]. The rate constant of ultra-violet photo-oxidation of the

organic molecule carbofuran using peroxodisulfate was reported as 0.98 min-1 by Chu

et al[53], while that for the photo-oxidation of oxalic acid using ozone was reported as

0.53 min-1 by Garoma and Gurol[54]. The kinetics of the oxidation process directly

determines the residence time necessary for digestion prior to detection. Since the

phosphomolybdenum blue reaction is relatively fast, it is the digestion step that

controls the sample throughput. The kinetics of ozone is comparable to that of

peroxodisulfate for similar organic molecules, which suggests that it may find

application for in-line digestion of total phosphorus. Dissolved ozone has the

additional advantage of being a “reagentless” digestion, in that aqueous ozone can be

generated by merging a stream of dry air that has been passed through a high voltage

electrical discharge with a stream of water in a flow analysis manifold.


48

2.1.5 Flow analysis methods for the in situ determination of phosphorus

Flow injection analysis has a been a commonly used method for measuring

phosphorus concentration in water samples since its development by Ruzicka and

Hansen in 1975[55]. Whilst extensive research has been devoted to developing

methods for analyzing both filterable reactive phosphorus and total phosphorus within

the laboratory, there has been a comparatively small amount of research into

instrumentation suitable for in situ determination of total phosphorus[18].

Traditionally, manual sampling and laboratory-based methods have been used to

acquire information regarding phosphorus concentrations in aquatic systems.

However, these protocols are costly in terms of both time and money, and present the

risk of sample degradation. The use of instrumentation designed for in situ

measurement can overcome these handling issues, as well as providing near real-time

analysis which can afford the opportunity of building spatial and temporal phosphorus

maps of high resolution.

In 1987, Worsfold et al[24] discussed the need for an automated in situ device for the

measurement of filterable reactive phosphorus in natural waters. Such a system would

need to meet the typical criteria imposed on a laboratory based instrument and also

require long-term stability, low power consumption, the ability to be periodically

calibrated and corrected accordingly, and the capability to function for extended

periods of operator absence. Over the past twenty years, there have been several

devices developed for the on-site measurement of filterable reactive phosphorus;

including micro-FIA units[56, 57], a 12-V battery operated flow injection apparatus


49

for phosphorus measurement[58], and a multi-commutation reagent injection flow

analysis instrument[18].

The portable instrument for the measurement of FRP developed by Lyddy-Meaney et

al[18], using molybdenum blue reactive phosphorus detection, achieved a sample

throughput of 380 measurements per hour, as well as high sensitivity and precision.

The lower detection limit of 5 µgPL-1 afforded by this system is more than adequate

to determine the total phosphorus concentration (subsequent to digestion) of most

natural waters, which is typically in the range 10 – 100 µgPL-1. The chromogenic

reagents were pressurised using inert gas prior to introduction by miniature solenoid

valves. Detection of phosphomolybdenum blue was achieved using a high sensitivity

multi-reflective cell coupled with a solid state light emitting diode and photodiode

detector[59]. The durable and economical nature of these devices is ideal for portable

field instrumentation, and provide a means of reducing reagent consumption and

waste generation.

While research has been undertaken toward developing portable field instrumentation

for in situ filterable reactive phosphorus measurement, comparatively little effort has

been directed to develop in situ measurement total phosphorus measurement

techniques. Determining total phosphorus has all the complications of measuring

molybdenum blue reactive phosphate, with the additional requirement of

mineralisation of all phosphorus containing species in a heterogeneous, unfiltered

sample.


50

Research objectives:

This chapter describes the design, construction and evaluation of a portable multi-

commutation reagent injection flow analysis field instrument for the measurement of

total phosphorus in natural waters, according to the following objectives:

• Development of a rapid, precise, sensitive and accurate portable flow analysis

method for the determination of total phosphorus in natural waters. This

method will be developed with in situ underway monitoring in mind, where

maximisation of reagent and power efficiency, and reagent longevity are

essential

• To investigate the viability of substituting dissolved ozone for the commonly

used peroxodisulfate oxidant

• Optimisation of method efficiency: design of thermal and photo reaction

chambers, altering digestion conditions to obtain optimum recovery of

refractory compounds at the fastest rate possible

• To test the developed and optimised method in a variety of field situations (in

fresh, estuarine and marine waters) to determine its suitability for in situ

deployment and application in nutrient mapping of natural waters.


51

2.2 Experimental

2.2.1 Reagents

Acidic molybdate reagent

In a 100 mL flask, 1.0 g of ammonium molybdate was dissolved in 50 mL ultrapure

water. Following dissolution, 3.5 mL of concentrated sulfuric acid was added and the

flask filled to 100 mL using ultrapure water.

Acidic tin(II) chloride-hydrazine reducing reagent

In a 100 mL flask, 0.2 g of hydrazine sulfate and 0.02 g of tin(II) chloride was

dissolved in 50 mL of ultrapure water. Following dissolution, 2.8 mL of concentrated

sulfuric acid was added and the flask filled to 100 mL using ultrapure water.

Acidic peroxodisulfate oxidant

In a 500 mL flask, 20.0 g of potassium peroxodisulfate was dissolved in 400 mL of

ultrapure water. Following dissolution, 0.7 mL of concentrated sulfuric acid was

added to the mixture, and the solution made to 500 mL using ultra pure water.

Alkaline peroxodisulfate oxidant

In a 500 mL flask, 20.0 g of potassium peroxodisulfate was dissolved in 400 mL of

ultrapure water. Following dissolution, 3.8 g of di-sodium tetraborate (Borax) was

added to the mixture, and the solution made to 500 mL using ultra pure water. The

final Borax concentration was 0.02 M.


52

Artificial seawater

1000 mL of artificial seawater was prepared as per the method described by Kester et

al[60].

Reactive phosphorus standards

In a 1000 mL flask, 0.4394 g of potassium dihydrogen phosphate was dissolved in

1000 mL ultrapure water to make a 100 mgPL-1 stock solution. This solution was

refrigerated and diluted as appropriate.

Organic phosphorus standards

In a 1000 mL flask, 0.3877 g of myo-inositol hexakisphosphate magnesium and

potassium salt (also called phytic acid) was dissolved in 1000 mL ultrapure water to

make a 100 mgPL-1 stock solution. This solution was refrigerated and diluted as

appropriate. Phytic acid was chosen because of its refractory nature.

Condensed phosphorus standards

In a 1000 mL flask, 0.3960 g of sodium tripolyphosphate was dissolved in 1000 mL

ultrapure water to make a 100 mgPL-1 stock solution. This solution was refrigerated

and diluted as appropriate.

Collection of natural water samples

All samples for TP determination were collected unfiltered and stored on ice during

transport. A Horiba probe was used to perform in situ measurements of pH, salinity,

turbidity, dissolved oxygen concentration, and temperature. The samples were stored

frozen until measured.


53

2.2.2 Instrumentation

Figure 2.1 A schematic representing the total phosphorus analyser. The digestion module and multi-commutation reagent injection analyser are indicated. Volumes and flow-rates are listed. Where R1, R2 = acidic molybdate reagent, acidic tin(II) chloride-hydrazine reagent respectively.

Sampler and digestion unit

A sampling and digestion module was used to handle all sample treatment operations

(Figure 2.1) including digestion, debubbling and filtration. Sample was collected

using a Masterflex peristaltic pump (model 7518-00) at 100 mLmin-1 using 33 mm i.d.

Masterflex silicon tubing. At this flow rate and tubing diameter there was no evidence

of settling or accumulation of particulate matter in any section of the sampling

system. Two Instech miniature peristaltic pumps (model P625) were used; one to

pump sample from the feed of the Masterflex pump at 2 mLmin-1, and another to

merge the sample stream with 2 mLmin-1 of acidic peroxodisulfate reagent or

UV Reactor Heat, 80 oC

Digestion reagent, 2 mLmin-1

Sample in, 2mLmin-1

Waste

Hollow-fibre filter, 300 µL

Debubbler

Waste, 0.8 mLmin-1

To analyser, 1.8 mLmin-1

2000 mm, 0.8 mm i.d. 1000 µL

600 mm, 0.5 mm i.d. 120 µL

R2 R1

Solenoid Valves

600 mm, 0.5 mm i.d.

LED 660 nm

Waste Peristaltic

Pump

Digestion Module

Analyser


54

dissolved ozone. The mixed sample-digestion reagent stream passed through a UV

photo-reactor consisting of a 12 W UV lamp (λmax = 254 nm) wound with 2000 mm

of 0.8 mm i.d. Teflon® tubing, and then through a 600 mm length of 0.5 mm i.d.

Teflon® tubing maintained at 80 oC by a 10 W heater. The digested sample was

filtered using a hollow fibre cross flow filter constructed of a single 100 mm length of

micro-porous polypropylene tubing (Accurel S6/2, 0.2 µm pore size, 1.8 mm i.d.,

Enka) supported internally by a multiply perforated piece of 0.5mm i.d. Teflon®

tubing. The hollow fibre was housed in a Perspex block (20 x 20 x 105 mm) inside a

chamber with a 2.5 mm i.d. bore, with both ends being sealed by glue. The digested

sample was introduced through a port in one end of the block. A length of 0.3 mm i.d

tubing connected to an exit port was used to increase the trans-membrane pressure

differential to enhance the flow rate through the membrane. The harsh acidic

oxidising conditions of the digested stream effectively prevent any particulate build-

up on the surface of the polypropylene tubing. The polypropylene tubular membrane

has an operating lifetime of approximately one week. The filtered digestate was then

passed through a debubbler, from where it was either drained to waste or was pumped

into the flow injection analyser. Automation of the sampler unit’s functions was

achieved using a USB-1608FS Measurement Computing™ A-D DAQ board

(National Instruments), interfaced to a personal computer running a LabView (v. 8.5)

control and data acquisition program (National Instruments). Figures 2.2 and 2.3

provide a pictorial representation of the digestion module.


55

Figure 2.2 A labeled picture of the digestion module.

Figure 2.3 A zoomed in and labeled picture of the digestion module.

UV Photo-reactor

Hollow-fibre Filter

Masterflex Pump

Heater Unit

Debubbler

Instech Pump

UV Photo-reactor Heater Unit

Hollow-fibre Filter

Debubbler


56

Reagent injection flow analyser

A multicommutation reagent injection flow analyser was used for the determination of

orthophosphate produced by in-line digestion (Figure 2.1). The sample stream is

driven by a peristaltic pump at 1.8 mLmin-1. Two solenoid valves are used to

introduce an acidic molybdate chromogenic reagent and an acidic tin(II) chloride-

hydrazine reducing agent from reagent storage chambers (6 mL volume) pressurised

with nitrogen gas at 50 kPa. 10 µL of each reagent is used per determination. The

sample and reagents were mixed using a 600 mm serpentine coil constructed from 0.5

mm i.d. Teflon® tubing threaded through a plastic support plate with holes drilled 0.3

mm apart to form a square grid pattern. A multi-reflective flow cell and red light

emitting diode (λmax = 660 nm) as described in Ellis et al[59] were used to detect

absorbance of the phosphomolybdenum blue generated. Automation of the analysers

functions was achieved using a USB-1608FS Measurement Computing™ A-D DAQ

board (National Instruments), interfaced to a personal computer running a LabView

(v. 8.5) control and data acquisition program (National Instruments).

Ozone Generator

Figure 2.4 A schematic of the in-house constructed ozone generator.

12 V

15 KHz Oscillator

Transformer

20 KV AC

Dry air in/ozonated air out

Glass tubing


57

An ozone generator utilising a dielectric corona discharge was constructed in-house,

as shown in Figure 2.4. A 12 V current is passed through a 15 KHz oscillator prior to

the voltage being increased to 20 KV by a TV flyback transformer that matches the

resonant frequency of the oscillating current. The transformed current was connected

to a 150 mm length of wire inside glass tubing, causing multiple dielectric corona

discharges and the ionisation of gaseous particles. A flow of dry air from a

compressed air cylinder passes through the dielectric discharge chamber, where the

oxygen was exposed to the 20 KV. The ozonated air was drawn from the opposite end

of the chamber using a peristaltic pump and merged with a stream of cold water; and

later used as a digestion reagent (Figure 2.1).

Sampling and analysis protocols for shipboard operation

The total phosphorus flow analyser was deployed aboard the SV Pelican 1 during

January in 2010 for the purpose of shipboard analysis of total phosphorus

concentration in natural waters around Port Philip and Westernport embayments and

Bass Strait, in Victoria, SE Australia. The analyser was calibrated every morning

using standards that were stored at less than 4 oC. A reference material was also

measured each morning to check the calibration accuracy. An organic phosphorus

standard of phytic acid, which is known to be recalcitrant, was used to check whether

100 % conversion was achieved in the digestion process. Recalibration using two

standards and the standard reference material was also done at the end of the day.

Sample was continuously pumped from a water intake in the hull at a depth of

approximately 0.5 m. A disc of 100 µm nylon mesh held in a 47 mm membrane filter

holder was used to prefilter the sample feed. Since the particulate phosphorus content

is inversely related to particle size[61], phosphorus contained in the particulate


58

fraction above 100 µm was deemed to be insignificant; and given the risk of blockage

in the FIA system these particles pose, this pretreatment was considered essential for

reliable operation. There was an estimated two minute delay between the sample

collection at the intake and the water feed at the sampler unit, as well as an additional

two minute residence time in the digestion module and analyser, giving an overall

four minute offset between sample collection and measurement, which equates to a

760 m spatial offset at the average cruise speed of the SV Pelican 1 (11.4 kmh-1).

Samples were collected at regular intervals after the sampler intake and immediately

frozen and retained for comparison with in situ measurements.


59

2.3 Results and discussion

As noted in the introduction to this chapter, a field instrument must meet the

performance requirements of a laboratory based technique (sensitivity, accuracy,

precision, selectivity) with the additional requirements of low power and reagent

consumption, long-term stability, the capacity for lengthy unattended operation and

calibration correction.

2.3.1 Suppression of silicomolybdenum blue interference in total phosphorus

measurements

It is also desirable for the method to have freedom from sample matrix effects. The

multi-reflective cell used in the orthophosphate analyser deployed in this method is

reported to significantly reduce refractive index effects[18]. The molybdenum blue

method which utilises a stannous chloride-hydrazine reducing agent is also reported to

lose sensitivity when the sample contains high concentrations of chloride ions[7],

which are prevalent in marine waters. Silicate[26] and arsenate[19] are also known

interfering species which react with acidic molybdate to form a heteropoly acid; with

silicon being present in high enough concentrations in natural waters to pose a

significant problem. Silicon interference is commonly reduced via decreasing the pH

of the sample stream below 1[26], which favors the formation of the

phosphomolybdate heteropoly acid. Figure 2.5 shows the relative peaks of a 250

µgPL-1 as orthophosphate standard and a 250 µgSiL-1 as silicic acid standard under

acidic reaction conditions (pH = 0.5) commonly used during total phosphorus


60

digestion. Figure 2.5 indicates that the signal unit response per microgram of silicon is

negligible in comparison to the response to phosphorus (approximately 1.6 % of the

phosphomolybdenum blue analytical response), which suggests that silicate

interference in natural waters will be nominal.

0

1500

3000

4500

6000

0 5 10 15 20 25

Time (s)

Dete

cto

r R

esp

on

se (

mV

)

250 ugP/L as orthophosphate 250 ugSi/L as silicic acid Ultrapure Water

Figure 2.5 A comparison of the analytical response of phosphomolybdenum and silicomolybdenum blue at 660 nm. Flow rate = 1.5 mLmin-1 40 gL-1 potassium peroxodisulfate in 0.025 M sulfuric acid; 3 mLmin-1 of standard/ultrapure water. The mixed acid-sample stream pH is 2.1 before reagent injection and 0.5 prior to reagent injection.

The use of antimonyl tartrate in conjunction with an ascorbic acid reductant has also

been reported to significantly reduce the formation of silicomolybdenum blue[62].

However, this method is highly temperature dependent and the formation of

phosphomolybdenum blue is slower than when using tin(II) chloride as a reductant.


61

2.3.2 Evaluation of dissolved ozone as a potential oxidant

The use of dissolved ozone generated in situ as an alternative to potassium

peroxodisulfate offers the potential of a reagent-generation system whereby

atmospheric oxygen could be used to generate the oxidising chemical species, which

would simplify field application.

Gaseous ozone was generated using a dielectric corona discharge (Figure 2.4), with

the ozonated air stream being merged with cold water before being introduced into the

digestion module (Figure 2.1). In order to test the effectiveness of dissolved ozone as

an oxidant, a stream of 100 µgPL-1 as phytic acid was merged in 1:1 ratio with a

stream of 0.18 mgO3L-1 (determined by batch ultra-violet spectroscopy), and passed

through a 3 m length of Teflon tubing wrapped around a mercury UV lamp (λmax =

254 nm). The mineralized zone was then measured in-line using the molybdenum blue

method and compared to a 100 µgPL-1 orthophosphate standard to determine the

percentage conversion. The results are shown in Figure 2.6.


62

0

400

800

1200

Mean

Peak H

eig

ht

(mV

)100 !gP/L as FRP

100 !gP/L as phytic acid -

no stop time


1 min stop time


2 min stop time

Figure 2.6 The relative mineralisation of phytic acid by dissolved ozone and photo-oxidation with 0, 1 and 2 minute stop times. Error bars are ± 1 σn-1 for n = 3.

The data indicate that the conversion of phytic acid to orthophosphate is incomplete

under the aforementioned conditions. While a longer stop time may produce a higher

yield of orthophosphate, the waiting time required outweighs the possible benefits of

using dissolved ozone. The rate at which dissolved ozone oxidises organic compounds

is proportional to the dissolved ozone concentration and the ultra-violet light

intensity[63]. Therefore, the likely reason for the poor result is the low concentration

of dissolved ozone (0.18 mgO3L-1) in comparison to the 40 gL-1 potassium

peroxodisulfate that was required to give rapid complete conversion[35]. The use of

dissolved ozone may be tenable if a means could be found to generate much higher

concentrations in the stream of air passing through the ozone generator, possibly via

using a stream of pure oxygen.


63

2.3.3 Optimisation of digestion conditions for total phosphorus measurement using

peroxodisulfate oxidant

Peroxodisulfate medium can effectively digest organic phosphorus compounds under

both acidic and alkaline conditions. However, there are several interferences in the

photo-oxidative alkaline digestion process in saline waters. For example; magnesium,

iron, aluminium and calcium ions may form stable photo-oxidant resistant complexes

with organic phosphorus compounds under alkaline conditions[48], scavenging of

hydroxide radicals may occur more readily under alkaline conditions than sulfate

radical scavenging under acidic conditions[45], and species such as nitrate, bromide

and chloride are strongly absorbed in the ultra-violet range (210 - 220 nm) and will

reduce available photons for photo-oxidation[46], although this will effect both acidic

and alkaline mediums. Aminot and Kerouel[46] found that increasing sample ionic

strength had a pronounced adverse effect on the photo-oxidative digestion of phytic

acid. In order to quantify the extent of interfering effects on photo-oxidative

performance, two standards of 200 µgPL-1 as phytic acid, one in artificial sea water

and another in ultra pure water, were measured using an alkaline peroxodisulfate

medium, buffered to pH 8.4 using 0.02 M borax buffer, and compared with an

equivalent concentration orthophosphate standard. The results are shown in Table 2.1.

Table 2.1 The effect of an artificial marine water matrix on the conversion of phytic acid to orthophosphate experienced by an alkaline peroxodisulfate digestion reagent.

Sample Measured P Conc. in µgPL-1 % Conversion

200 µgPL-1 phytic acid in ultrapure water 211 ± 6 105 ± 3

200 µgPL-1 phytic acid in artificial seawater 41 ± 2 20 ± 1


64

The results from this experiment indicate that radical scavenging and/or metal

complex formation has a significant impact on the oxidation efficiency of an alkaline

peroxodisulfate medium when measuring saline samples. An acidic medium was used

to overcome this as reported by Peat et al[48]. Using an acidic oxidant also simplifies

digestion, as an additional acidic step for the dissolution of particulate phosphorus and

hydrolysis of condensed phosphates is not required.

In order to measure the conversion efficiency of condensed phosphate species, a

stream of sulfuric acid of varying concentration was merged with a stream of 100

µgPL-1 as sodium tripolyphosphate standard (Figure 2.1). While increasing acid

strength will increase the rate of hydrolysis of condensed phosphates, it also has the

unwanted effect of suppressing the formation of the phosphomolybdenum blue. These

trends are shown in Figure 2.7.

0

25

50

75

100

0.0 0.5 1.0 1.5 2.0

Concentration of sulfuric acid (M)

% C

on

ve

rsio

n o

f

trip

oly

ph

os

ph

ate

to

ort

ho

ph

os

ph

ate

0

25

50

75

100

% S

en

sit

ivit

y s

up

pre

ss

ion

% Conversion of tripolyphosphate % Sensitivity suppression

Figure 2.7 The conversion of sodium tripolyphosphate to reactive phosphorus with varying sulfuric acid concentration, as well as the suppression of the molybdenum detection chemistry. Error bars are ± 1 σn-1 for n = 3.


65

Figure 2.7 indicates that while increasing the acid concentration does dramatically

increase the conversion efficiency of the condensed phosphates, there is an

accompanying sensitivity loss (98.4 % suppression of sensitivity) at the 2 M sulfuric

acid concentration that is required to give complete hydrolysis using continuous flow.

Condensed phosphate species have a half-life of hours in natural waters[64], and thus

are typically found only in very low concentrations. Therefore, the sacrifice in

sensitivity that occurs in order to achieve complete mineralisation of these species is

not worth the gain in conversion, except perhaps if the instrument was to be deployed

for the monitoring of waste waters, where condensed phosphates might be a

significant component of the total phosphorus concentration. For all subsequent work,

an operating concentration of 0.025 M sulfuric acid was chosen.

Oxidant concentrations of up to 40 gL-1 peroxodisulfate are reported in the literature

to be necessary to achieve 100 % conversion of refractory organic phosphorus

compounds by photo-oxidation[35]. It should be noted that 40 gL-1 potassium

peroxodisulfate also represents the practical solubility limit of this reagent under

typical operating conditions. The change in conversion efficiency with varying

concentration of peroxodisulfate was determined by measuring the conversion of a

100 µgPL-1 as phytic acid using 0, 10, 20, 30 and 40 gL-1 peroxodisulfate solutions in

0.025 M sulfuric acid. The results are shown in Figure 2.8.


66

0

25

50

75

100

0 10 20 30 40

Concentration of potassium peroxodisulfate (gL-1

)

% C

on

vers

ion

of

ph

yti

c a

cid

to

ort

ho

ph

osp

hate

Figure 2.8 The change in conversion efficiency with varying peroxodisulfate concentration. A 2000 mm length of 0.8 mm i.d. Teflon tubing wound around a medium pressure mercury UV lamp was used. Error bar are ± 1 σn-1 for n = 3.

Figure 2.8 indicates that 40 gL-1 peroxodisulfate solution is required to achieve 100 %

conversion for solutions of the refractory organic phosphorus compound phytic acid.

In addition, it is interesting to note that a solution of 0.025 M sulfuric acid will still

produce 40.5 % conversion of phytic acid to orthophosphate through hydrolysis alone

when the mixture is exposed to ultra-violet radiation and heated to 80 oC.

The primary mechanism by which photo-oxidation enhances the conversion efficiency

of peroxodisulfate is thought to be via the generation of hydroxyl and sulfate radicals,

as represented in Equations 2.4 and 2.5. Presumably, an increase in ultra-violet

irradiation would increase the number of radicals generated, and hence lead to

improved digestion efficiency. An increase in exposure can be achieved in three ways:

increasing the length of the photo-reactor coil, decreasing the wall thickness of the

photo-reactor coil tubing and thus decreasing absorption of ultra-violet light by the


67

tubing wall, or by extending the irradiation period through the introduction of a stop

flow time and therefore producing a higher concentration of radicals.

While the stop-flow procedure is more economical from a reagent and sample usage

stand-point, it also reduces the sample throughput and increases instrumental

complexity. With the initial design goals in mind, instrumental simplicity and rapid

throughput were chosen in preference to reduced reagent and sample consumption.

An increase in the photo-reactor tubing coil length to gain more UV exposure would

enable the benefit of continuous flow through the digestion module while maintaining

high digestion efficiency. The photo-reactor was constructed using tubing with a wall

thickness of 0.4 mm (0.8 mm i.d.), in preference to 0.55 mm wall thickness tubing

(0.5mm i.d.) in order to decrease the ultra-violet absorption by the tubing walls.

The minimum tubing length required to achieve 100 % conversion was determined by

measuring the conversion of a 200 µgPL-1 phytic acid standard mixed with 40 gL-1

peroxodisulfate in 0.025 M sulfuric acid over 500, 1000, 1500, and 2000 mm lengths

of reactor coil. The results are shown in Figure 2.9 below, and as expected, an

increase in tubing length leads to an increase in conversion efficiency, where 97 % (±

1.4 %) conversion is achieved for a coil length of 2000 mm.


68

0

25

50

75

100

500 1000 1500 2000

Photo-reactor tubing length (mm)

% C

on

vers

ion

of

ph

yti

c a

cid

to

ort

ho

ph

osp

hate

Figure 2.9 Oxidation of phytic acid with varying photo-reactor tubing length. Error bars are ± 1 σn-1 for n = 3.

2.3.4 Analytical figures of merit

As the digestion module produces a continuously flowing stream of mineralised

sample, the number of measurements that can be achieved within a given time-frame

is dependent upon the speed of the colorimetric reaction. Figure 2.10 below illustrates

that at a flow rate of 1.8 mLmin-1 a peak may be collected every 13 seconds.

Throughput may be increased by increasing the analyser flow rate, but this will

prevent the phosphomolybdenum blue formation from proceeding to completion[18],

causing a loss of sensitivity. Additionally, the reagent consumption required to

achieve a throughput of one sample every 13 seconds is excessive in comparison to

the extra resolution this throughput may offer during field operation. For these

reasons, an operating cycle of one sample per 31 seconds was found to be optimal.


69

0

1000

2000

3000

0 40 80 120 160

Time elapsed (s)

Dete

cto

r O

utp

ut

(mV

)

Figure 2.10 Replicate peak responses for a blank, 50, 100 and 200 µgPL-1 orthophosphate standards. The blank is higher than normally observed for orthophosphate measurements because the 40 gL-1 peroxodisulfate reagent contains some phosphorus contamination.

The analytical figures of merit, derived from the data shown in Figure 2.10 are listed

in Table 2.2.

Table 2.2 The analytical figures of merit for the flow analysis system. The limit of detection is determined by the linear regression method described by Miller and Miller[65]. Sensitivity 10.41 mV/µgPL-1

Precision (%RSD on 100 µgPL-1 phytic acid) 4.6 % (n = 10)

Throughput 115 measurements per hour Limit of Detection (99% conf. limit) 1 µgPL-1

Limit of Quantification (10σblank) 13 µgPL-1 Linearity (over the calibration range 0–200 µgPL-1) R2 = 0.9998

The proposed method displayed excellent sensitivity and linearity over the 0 – 200

µgPL-1 calibration range, which resulted in a detection limit of 1 µgPL-1. This


70

detection limit is adequate for marine coastal and most pristine fresh waters where the

total phosphorus concentration may typically be less than 10 µgPL-1.

A precision of 4.6 %RSD was obtained when measuring a 100 µgPL-1 phytic acid

standard (n = 10), which indicates that the conversion of phytic acid to

orthophosphate is quite repeatable. The repeatability could be further improved by use

of a peristaltic pump with more rollers (e.g. 6 - 8) rather than the three-roller pumps

employed in these experiments to merge the sample and peroxodisulfate streams. The

photometric sensitivity of the phosphomolybdenum blue method is highly dependent

on the acid concentration, and any variability in the sample and digestion reagent

mixing ratio resulting from the use of three roller pumps will cause fluctuations in

observed peak heights.

2.3.5 Laboratory evaluation of the optimised technique

There is the potential for sample matrix effects of some natural waters to reduce the

conversion efficiency of the digestion method, in addition to interfering with or

suppressing the colorimetric detection chemistry. Species such as carbonate may act

as radical scavengers, and cations found in natural waters may reduce the oxidative

effectiveness of the digestion method by forming stable phosphate complexes[48].

Silicate, arsenate and germanate will also form heteropoly acids in the presence of

acidic molybdate[26], leading to an overestimation of total phosphorus. Chloride ions

may also suppress the formation of phosphomolybdenum blue[18] leading to an

underestimation of total phosphorus. If the injected reagents do not mix completely


71

with the sample, there is also the potential for refractive index effects to occur for

saline samples.

In order to test the tolerance of the developed total phosphorus flow analysis method

to various sample matrices, comparative analysis was performed on nine samples

collected in Port Philip Bay and the Yarra River estuary. These samples exhibited the

wide range of salinity and turbidity that might be expected from such a marine-

estuarine-freshwater system, as seen in Table 2.3.

Table 2.3 Properties of the water samples collected, as measured in situ. The salinity column indicates if the samples are marine (M), estuarine (E), or freshwater (F). Samples 1-2 are from Port Philip Bay, and 3 - 9 from the Yarra River estuary.

Sample location Salinity pH Temp. (oC) DO (mgO2L-1) Turbidity (NTU)

1. Williamstown Jetty 34.7 (M) 7.9 17.6 8.4 4 2. Williamstown

Foreshore 34.9 (M) 8.0 17.9 9.4 6

3. Westgate Bridge 29.4 (E) 7.9 18.3 7.9 4 4. Federation Square 10.2 (E) 7.4 19.8 5.5 5

5. Morell Bridge 10.8 (E) 7.4 19.5 5.9 7 6. Herring Island 7.1 (E) 7.2 19.0 5.7 8

7. St Kevin’s College Boathouse 5.1 (E) 7.2 19.6 6.3 11

8. Hawthorn Bridge 0.9 (E) 7.4 20.0 6.6 23 9. Fairfield Park 0.0 (F) 7.4 19.7 5.3 22

These samples are varied in salinity (0 - 35) and turbidity (4 - 22 NTU) sufficiently to

test the methods tolerance for wide variations in sample matrices. The samples were

collected using the flow analysis system without any filtering or pretreatment. The

digestion module and analyser required a flush time of 90 seconds to clear the

previous sample completely, after which a replicate measurement was performed

every 30 seconds. Thus, one sample can be analyzed in triplicate every 180 seconds,

giving an overall sampling rate of 20 samples in triplicate per hour. The


72

measurements obtained by the proposed flow analysis method were validated against

an established total phosphorus method, involving autoclave digestion with

peroxodisulfate followed by spectrophotometric detection[66]. As shown in Figures

2.11 and 2.12, the results showed a good degree of agreement, with no apparent

discrepancies for saline and less saline samples, indicating that chloride ion or

digestion interference from cations were negligible. A Wilcoxon signed rank test

(Ptwo-tail = 0.045, n = 9) indicates that there is possibly some discrepancy between the

two data sets. The developed method is also capable of handling samples with a high

turbidity without exhibiting any blockages or spectral interference from fine

particulate organic matter.

0

50

100

150

1 2 3 4 5 6 7 8 9

Sample number

To

tal P

ho

sp

ho

rus _

gP

L-1

Flow analysis

method

Comparative

method

Figure 2.11 A bar chart comparing the total phosphorus concentration as determined by the flow analysis method and the comparative method[66]. Sample details are found in Table 2.3. Error bars are ± 1 σn-1 for n = 3. .


73

y = (0.90 ± 0.07)x

+ (14.36 ± 6.44)

R2 = 0.9627

50

75

100

125

150

50 75 100 125 150

Comparative method total phosphorus (_gPL-1

)

Flo

w a

na

lys

is t

ota

l p

ho

sp

ho

rus

(_

gP

L-1

)

Figure 2.12 A comparative line chart indicating the approximate 10 % bias towards the proposed flow analysis method. Error bars are ± 1 σn-1 for n= 3.

These results demonstrate that acidic peroxodisulfate digestion is markedly more

effective in handling digestion of organic phosphorus compounds in saline waters

than an alkaline medium (Table 2.1). However, both the Wilcoxon signed rank test

and the comparative function (Figure 2.12) indicate that the proposed method slightly

overestimates the sample total phosphorus concentration to the comparative method.

As the overestimation is reasonably consistent, as indicated by the linearity of the

comparative function (Figure 2.12), this is most likely due to errors in calibration

rather than any effect of sample matrix.

2.3.6 Instrumental and method stability

There are several design factors that determine the long-term viability and stability of

an instrumental method; namely reagent stability, reagent economy, power


74

consumption, and instrumental durability. In order to optimise reagent stability and

economy, the method must utilise reagents that are capable of both long-term storage

without degradation and the ability to produce highly sensitive analytical responses

while using minimal volumes. The use of reagent storage chambers pressurised with

inert gases such as argon, helium and nitrogen can be used as a means of delaying

reagent decomposition. It is also desirable that the inert gas used be of low solubility,

thus minimising any out-gassing or bubble formation upon reagent injection. This is

particularly relevant to phosphorus determination, as tin(II) chloride is thought to

readily undergo oxidation despite being stabilised by hydrazine sulfate. Lyddey-

Meaney et al[18] determined that the acidic molybdate and acidic tin(II) chloride-

hydrazine sulfate reagent with tin(II) chloride showed no signs of degradation when

stored under inert gas over a period of two weeks. Gas pressurised reagent injection is

also very efficient, as the potential energy of the gas is used to force the small reagent

volumes into the liquid stream rather than electrical energy, as used by a peristaltic or

syringe pump.

If small volumes of reagents are used (less than 20 µL), then a small sample-reagent

zone will be formed, which means that the sensitivity of the method will be

particularly prone to decline as dispersion of the sample-reagent zone increases. In

order to limit sample zone dispersion, manifold volume and residence time must be

kept to a minimum. The multi-reflective cell deployed in the reagent injection

analyser has a significantly higher pathlength to volume ratio than a standard 10 mm

pathlength z-configuration flow cell[59]. Dispersion may also be further reduced by

placing the detector close to the injection point. However, if the chromogenic reaction

is not instantaneous this may also reduce sensitivity because of incomplete


75

chromogenic reactions. For this reason, a tin(II) chloride-hydrazine sulfate reductant

is preferred over ascorbic acid reductant[2], because of its faster reduction kinetics. A

600 mm 0.5 mm i.d. Teflon® mixing coil placed between the injection valves and the

detector was found to offer the ideal balance between dispersion and reaction

completion. Use of small reagent injection volumes may also limit the upper end of

the dynamic linear range, and may lead to diminished sensitivity, as the presence of

larger quantities of analyte will effectively exhaust the limited concentration of

reagent present in the mixed sample-reagent zone.

The instrument components must also be sufficiently robust to withstand the heated,

acidic and strongly oxidising reaction mixture, in addition to the rigors of field use.

Inert Teflon® tubing and fittings were used to avoid any corrosion of the operating

manifold. The analyser and sampling unit are designed to operate from a 12 V DC

source, drawing approximately 36 W with all components operating. The ultra-violet

germicidal lamp was only activated when necessary, in order to minimise operating

power requirements.

In order to test the stability of the flow analysis total phosphorus method, a two-week

test was undertaken. The instrument was programmed to perform triplicate

measurements of a test sample every 60 minutes. This test sample was collected from

the Herring Island site in the Yarra River estuary, and had similar properties to the

sample collected from the same site in earlier experiments (Table 2.3). The same

sample was used for repeated measurements over the two-week time period. Every

24-hour interval, a portion of the sample was collected and frozen for later

comparison with an established method[66], and a 100 µgPL-1 orthophosphate


76

standard was measured to correct for any signal drift. Figure 2.13 displays the results

of the two-week test.

0

50

100

150

0 50 100 150 200 250 300 350

Hours

To

tal p

ho

sp

ho

rus (!

gP

L-1

)

[P] vs Time, Flow analysis method Comparative Method

Figure 2.13 Automated determination of total phosphorus by the flow analysis method over a two week period. Results from a comparative method[66] are also included.

The instrument operated successfully over the 336 hour test duration, excluding an

event between 110 and 120 hours during which the analyser lost gas pressure due to a

leak, and as such no reagents were injected into the sample stream causing only a

baseline peak to be obtained. The measurements from the comparative method

indicate good agreement with the flow analysis method. A Wilcoxon signed rank test

(Ptwo-tail = 0.035, n = 14) indicates that there is some bias in the measurements that has

not arisen by chance. This discrepancy is clear in Figure 2.13, where there is a

noticable difference between the values derived from the flow analysis and

comparative methods after the 250 hour mark. This is most likely due to two factors:

it was discovered subsequently that the aqueous peroxodisulfate was undergoing


77

degradation when in solution, forming oxygen and sulfuric acid upon

decomposition[67], according to Equation 2.8:

S2O82-

(aq) + H2O(aq) � 1/2O2(g) + 2SO42-

(aq) + 2H+(aq) (2.8)

This can cause a loss of analytical response both by further acidifying the sample

stream (causing suppression of the formation of phosphomolybdenum blue) and by

decreasing oxidant concentration. The former is most likely in this case, as any

suppression of the analytical response due to increased acidification should be also

observed in the standard used to correct for instrumental drift. The second possibility

is that the single stage regulator used to control the gas pressure, which propels a

volume of chromogenic reagents into the digested stream, may have undergone some

pressure drift, causing successively smaller volumes of reagents to be introduced.

The combined flow-rate of the sample and oxidising agent was observed to drop from

3.5 mLmin-1 to 2.9 mLmin-1 over the 356 hour course of the experiment. The loss of

flow rate was partially caused by wear to the pump tubing and very minor

accumulation of particulate matter on the surface of the hollow-fibre filter. Despite

partial blockage over the two-week long test, these results indicate that the hollow-

fibre filter is reliable over a long-term timeframe.

A reagent volume of 5.0 mL of acidic molybdate and acidic hydrazine sulfate reagents

was sufficient for 120 hours of operation, or 480 injections. The injection volume for

each reagent was approximately 10 µL, or 40 µL per triplicate measurements,

including flushes. The system generates waste of 4 mLmin-1 which consists of sample

mixed with acidic peroxodisulfate agent in a 1:1 ratio.


78

To further investigate the problems encountered during the two week trial, a second

one week trial was undertaken using a larger volume of sample collected in the same

location. The results are displayed in Figure 2.14 below.

0

25

50

75

100

125

0 24 48 72 96 120 144 168

Time (hours)

To

tal p

ho

sp

ho

rus (

_g

L-1

)

0

25

50

75

100

125

% c

on

vers

ion

Flow analysis method

Comparative method

% conversion

Figure 2.14 The measured phosphorus concentration of a sample over 168 hours as determined by the proposed flow analysis method. The results of a validation method are also shown, as well as the percentage conversion of a phytic acid standard measured every 24 hours. The dashed line represents 100 % conversion of the phytic acid standard. For this experiment, the acidic peroxodisulfate solution was renewed on a daily basis

and daily checks of conversion efficiency were performed using a 100 µgPL-1

standard of phytic acid. The internal gas pressure in the reagent chambers was also

monitored regularly. The relative standard deviation for the analytical response of the

100 µgPL-1 orthophosphate standard was 16 % (n = 21, 3 measurements per day) over

the 168 hour period, which indicates that instrumental drift was minimal. The one

week trial data showed an improved degree of agreement between the flow analysis


79

method and the comparative method, except for one point at 96 hours, where the

phytic acid standard showed that the conversion efficiency had dropped to 62 %. This

is most likely due to a faulty batch of the peroxodisulfate reagent, as greater than 90

% conversion was achieved from the next day following replacement of the oxidant.

A Wilcoxon signed rank test (Ptwo-tail = 0.81, n = 7) indicates no overall bias between

the comparative method and the proposed flow analysis method. Figure 2.14 indicates

that the colorimetric reagents were stable over the one week period and capable of

producing precise results over an extended period of time. However, installation of a

two or three stage gas regulator to provide superior long-term pressure moderation

would go a long way towards increasing instrumental precision.

2.3.7 Results of continuous in situ total phosphorus measurement during the Two

Bays study

The flow analysis total phosphorus system was deployed aboard the SV Pelican 1

during a two week scientific study of Port Philip and Westernport Bays and Bass

Strait (the Two Bays program) in Victoria, SE Australia, during January 2010. The

main objective of this deployment was to perform real time, in situ total phosphorus

mapping of both embayments and Bass Strait.


80

-38.6

-38.5

-38.4

-38.3

-38.2

-38.1

-38

-37.9

-37.8

144.3 144.5 144.7 144.9 145.1 145.3 145.5

Latitude (o)

Lo

ng

itu

de

(o)

80 to 110_gP\L 60 to 80_gP\L 40 to 60_gP\L 20 to 40_gP\L 5 to 20_gP\L

Figure 2.15 A map indicating the total phosphorus concentration (5-110µgPL-1) as determined in situ at locations recorded using a GPS unit. The three legs of the journey are; (A) Docklands to Rye, (B) Portarlington to Geelong and then to Williamstown, (C) Queenscliff to Hastings and then clockwise around French Island.

Figure 2.15 shows 1236 measurements of total phosphorus concentration collected

over a 25 hour period, over a distance of approximately 285 kilometres. In total 2499

points were recorded; but of these, 542 (22 %) were lost due to a malfunction in GPS

logging and another 721 (29 %) were discarded due to bubbles causing distortion of

the peaks. Most notably, data for the entire leg between Hastings and the northwestern

point of French Island, and between Queenscliff and Portarlington could not be used

for mapping purposes due to GPS failure. While the data loss due to the GPS fault can

be easily rectified during future field use, the lost peaks owing to bubble evolution

continue to be a problem. Even allowing for this data loss, a spatial resolution of 1

Leg A

Leg C

Leg B


81

measurement every 230 metres was achieved for an average cruise speed of 11.4

kilometres per hour.

Samples were collected and frozen at intervals throughout the cruise to validate the in

situ measurements taken by the flow analysis method. Figure 2.16 shows that there is

strong agreement between the two sets of values, with the slope of the comparative

function being 1.038 ± 0.036.

y = (1.0375 ± 0.036)x

- (0.183 ± 4.179)

R2 = 0.9776

0

25

50

75

100

0 25 50 75 100

Total phosphorus (!gL-1

) by comparative method

To

tal p

ho

sp

ho

rus (!

gL

-1) b

y fl

ow

an

aly

sis

meth

od

Figure 2.16 A comparative chart indicating strong agreement with the comparative method[66] and continuous flow in situ measurements. The error in the gradient and y-intercept are shown within the brackets. The dashed line represents a 1:1 agreement.

The scatter about the regression line suggests that there is no overall clear bias

between the total phosphorus concentration determined in situ and that determined

later by the comparative method[66]. A Wilcoxon signed rank test (Ptwo-tail = 0.78, n =

21) indicates no overall bias between the comparative method and the in situ

measurements by the proposed flow analysis method. It is possible that the collected

samples may have undergone some degradation prior to analysis because of the very


82

hot conditions that prevailed during the SV Pelican 1 cruise (temperatures of 45 oC on

deck) and the extended time required to completely freeze the samples (up to 24

hours). In addition, each in situ measurement was determined from only a single

analysis, which increases uncertainty as the average relative standard deviation of the

flow analysis method has been determined to be up to 4.6 % (Table 2.2). Despite

these limitations the agreement between the two methods is excellent.

2.3.8 Interpretation of the total phosphorus data obtained during the Two Bays cruise

Figure 2.15 shows that the total phosphorus concentrations in Port Philip Bay are

much higher than in Bass Strait and Westernport Bay. This is caused by two factors:

waters in Port Philip Bay experience a high residence time (estimated to be 6

months[68] to 12 months[69]) because of the narrow channel between the Port Philip

heads, and thus mixing with the oligotrophic ocean waters of Bass Strait is limited,

and the anthropogenic phosphorus inputs from the industrial activity, water treatment

facilities and the highly populated coastal regions of Port Philip Bay.

Nutrient inputs to Port Phillip Bay are the Yarra and Patterson Rivers, the Mordialloc

Creek drain and other minor creeks, and the atmosphere in the case nitrogen. The

waste treatment plant located at Werribee on the northwestern coast of the bay is also

a major anthropogenic source of nutrients[69]. While nitrogen concentrations are

generally low in the bay, due to denitrification in the sediments[70], phosphorus

concentrations are predominantly regulated by ocean transport, and are thus usually

higher due to incremental build up over time.


83

The 1992 ANZECC guidelines for total phosphorus concentrations in coastal waters

are 1 – 10 µgPL-1 as phosphorus-phosphate and 5 – 15 µgPL-1 as total phosphorus in

estuaries and embayments[11]. While these guidelines have subsequently been

changed to a more “risk based” approach over the past decade (ANZECC, 2000)[71],

the aforementioned guidelines remain useful for assessing apparent risk to aquatic

system health.

0

25

50

75

100

125

8:52 10:04 11:16 12:28 13:40 14:52 16:04

Time

To

tal p

ho

sp

ho

rus (!

gL

-1)

Flow analysis method Comparative method

Yarra River

Mordialloc Creek

Patterson River

Safety Beach Marina

Figure 2.17 Total phosphorus concentration as determined in situ and plotted against time for 11-Jan-2010. Validation data is plotted along with the in situ measurements for samples collected throughout the cruise. The cruise began at Docklands, and then proceeded down the coast of the Mornington Peninsula inside Port Philip Bay (Leg A in Fig 2.15). Four points of interested are indicated on the chart.

The total phosphorus concentration in the Yarra Estuary is typically higher than that

of Port Philip Bay. Along the western coast of the Mornington Peninsula are three

inputs of interest: Mordialloc Creek drain which can be seen prominently in Figure

2.17, the Patterson River, and the Safety Beach Marina/Tassell’s Creek. These inputs

are labeled in Figure 2.18 at the corresponding time to which the SV Pelican 1 sailed

past them, using the recorded GPS values. There is a clear observable increase in


84

phosphorus concentrations in these areas, most notably the Mordialloc Creek, which

drains both industrial and agricultural lands.

0

25

50

75

100

125

8:09 9:07 10:04 11:02 12:00 12:57

Time

To

tal p

ho

sp

ho

rus (!

gL

-1)


Waste Treatment

Plant, Werribee Laverton Creek

Figure 2.18 Total phosphorus concentration as determined in situ and plotted against time for 23-Jan-2010. Validation data is plotted along with the in situ measurements for samples collected throughout the cruise. The cruise began at Geelong, and then proceeded down the northwestern coast of Port Philip Bay to Williamstown (Leg B in Figure 2.15). Two points of interested are indicated on the chart, the waste treatment plant located near Werribee and Laverton Creek.

As can be seen in Figures 2.15 and 2.18, total phosphorus concentrations increase

significantly upon approach into Corio Bay. This is most likely due to the increased

anthropogenic and industrial activity in this area, as well as mixing with high nutrient

waters from the waste treatment plant at Werribee. The waste treatment plant is

located along the coastline north of Corio Bay, and the treated waster is released

directly into Port Philip Bay from several pipelines. Total phosphorus levels in the

water adjacent the treatment plant reached concentrations of up to 40 µgPL-1 higher

than the ambient waters. An increase in phosphorus concentrations was also detected

near Laverton Creek, a waterway that passes through urban and intensely industrial

areas before discharging into a coastal wetland.


85

0

10

20

30

40

50

10:04 11:19 12:34 13:49 15:04 16:19

Time

To

tal p

ho

sp

ho

rus (

_g

L-1

)


Boag's Rocks

Outfall,

Gunnamatta

Figure 2.19 Total phosphorus concentration as determined in situ and plotted against time for 12-Jan-2010. Validation data is plotted along with the in situ measurements for samples collected throughout the cruise. The cruise began at Queenscliff, through the heads, and then proceeded down the southern coast of the Mornington Peninsula/Bass Strait, into Western Port Bay to Hastings (Leg C in Figure 2.15). The Boag’s Rocks outfall near Gunnamatta is indicated on the chart.

In comparison to Port Philip Bay, the waters of Bass Strait are relatively pristine,

which is supported by the data in both Figure 2.15 and 2.19. The Strait is well mixed

internally and has free exchange with the oligotrophic Southern and Tasman

oceans[69], and hence has low concentrations of nutrients.

The Boag’s Rock outfall discharges treated sewage from the Eastern Treatment Plant

via a pipeline into Bass Strait. While the quantity of nutrients released from Boag’s

Rock is comparable to the treatment plant at Werribee, strong advective currents act

to transport nutrients away from the outfall and disperse the plume[69]; nonetheless

the potential for significant environmental impact still exists at this site. The

measurements in Figure 2.19 reflect these mitigating effects, as the low nutrient

concentration of the ambient waters and the currents keep the total phosphorus


86

concentration in the dispersed plume to a maximum of 34 µgPL-1, three times lower

than that of the waters adjacent the Werribee Treatment Plant.

The data in Figures 2.17-19 indicates the developed total phosphorus flow analyser

can be applied to resolve individual point sources of phosphorus input from various

urban creeks and sewage outfalls as measured in situ, in real-time, with a high degree

of accuracy, and with greater efficiency and cost effectiveness than that offered by

manual sampling. The 1778 valid measurements performed during this cruise over a

period of 5 days represent 10 - 20 days of conventional laboratory analysis (assuming

90 - 180 samples processed per day) and would cost approximately A$40,000. Given

that the instrument costs approximately A$10,000 and A$500 worth of reagents were

used in this exercise, the large savings in time and cost (even allowing for operator

time) combined with the high intensity and quality data, testifies to the benefits

offered by portable flow injection instrumentation.


87

2.4 Conclusion

The overarching aim of this work was to construct a total phosphorus analyser capable

of rapid, reliable, accurate measurements with the capacity for stand-alone field

operation for extended periods of time. This chapter reports that use of an in-line

photo-reactor and thermal decomposition unit, coupled with a reagent injection flow

analyser was successfully used for rapid and reliable determination of total

phosphorus in surface marine, estuarine and freshwaters.

It was demonstrated that 100 % mineralisation of total phosphorus to orthophosphate

in natural water samples can be achieved reliably and rapidly using an acidic

peroxodisulfate oxidising medium coupled with ultra-violet photo-oxidation. Phytic

acid, used as a surrogate for organic phosphorus species, was completely mineralised

to orthophosphate. However, complete hydrolysis of condensed phosphate species to

orthophosphate on a rapid timescale (<1 minute) remains difficult to achieve without

sacrificing a large degree of sensitivity (up to 98 %). The proposed flow analysis

procedure is free from sample matrix effects. Ozone was eliminated as an alternative

oxidant to peroxodisulfate due to the difficulty in generating high concentrations of

dissolved ozone in solution.

The phosphomolybdenum blue method of determining orthophosphate generated from

in-line digestion was found to be free of sample matrix effects and interference from

orthosilicate. This method provided excellent sensitivity and a lower detection limit of

1 µgPL-1, which is adequate for most natural waters. A sample throughput of 115

measurements per hour was achieved. Further improvement of the instrumental


88

repeatability could be attained through the use of better quality peristaltic pumps in

the digestion module.

The total phosphorus flow analyser has been shown to be capable of operation for a

period of up to one week in a stand-alone fashion, while taking a triplicate

measurement of sample once every hour. In order to extend this time or to perform

more measurements per hour, the volume of the reagent chambers would need to be

increased. In addition, the accuracy of the method has been found to be particularly

prone to fluctuations in gas pressure within the reagent chamber, and modification of

the existing equipment to include a two or three stage regulator is required to improve

the long-term stability of the analyser. While unattended operation of the system is

limited to around 96 hours using a combined peroxodisulfate and acid reagent, this

could feasibly be extended by storing these reagents separately and mixing them in-

line prior to digestion.

The developed instrument has been applied successfully in the field for the mapping

of total phosphorus in Port Philip and Western Port Bays, Victoria, SE Australia.

Samples taken concurrently for validation suggest a high degree of accuracy in the

measurements. 2499 measurements were recorded over the course of 25 hours and

approximately 285 kilometres; however, only 1236 were able to be used for mapping

purposes due to bubble evolution and a malfunction in GPS logging. Even with that

data loss, these analyses yielded a temporal resolution of 73 seconds, and a spatial

resolution of 230 metres at an average cruise speed of 11.4 kmh-1, which is more than

adequate to provide information on point-source inputs and produce real-time nutrient

mapping. While the GPS fault can be easily rectified for future cruises, the data loss


89

due to bubble evolution remains a confounding factor when measuring natural

samples in the field.


90

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oxidation of TCE at ambient temperatures. Chemosphere 66, 106–113.

46. Aminot, A., and Kerouel, R. (2001). An automated photo-oxidation method

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water. Marine Chemistry 76, 113–126.

47. Walsh, T.W. (1989). Total dissolved nitrogen in seawater: A new high

temperature combustion method and a comparison with photo-oxidation.

Marine Chemistry 26, 295-311.

48. Peat, D.M.W., McKelvie, I.D., Matthews, G.P., Haygarth, P.M., and

Worsfold, P.J. (1997). Rapid determination of dissolved organic phosphorus in

soil leachates and runoff waters by flow injection analysis with on-line photo-

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49. Rumhayati, B. (2007). In situ measurement of phosphorus species in overlying

and porewaters using the La(OH)3 diffusive gradient in thin films. PhD thesis,

Monash University, Clayton.

50. Ohashi, K., Yasu, K., Suzuki, C., and Yamamoto, K. (1977). A

spectrophotometric study of phosphomolybdenum blue formed by reaction of

phosphate with a mixutre of molybdenum(V) and molybdenum(VI) and

application to the spectrophotometric determination to small amounts of

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57. Motomizu, S., Oshima, M., and Ma, L. (1997). On-site analysis for

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phosphorus during rain events. Biogeochemistry 47, 87-109.

62. Kuroda, R., Ida, I., and Oguma, K. (1984). Determination of phosphorus in

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63. Kusakabe, K., Aso, S., Wada, T., Hayashi, J.-I., Morooka, S., and Isomura, K.

(1991). Destruction rate of volatile organochlorine compounds in water by

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64. Sawyer, C.N. (1952). Some new aspects of phosphates in relation to lake

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99

Chapter 3 – Design and construction of a total

internal reflective flow cell for use in flow

analysis

Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis

100

3.1 Introduction

Flow injection analysis is a versatile approach for the handling of various liquid

analyses[1]. The flexibility of flow injection techniques is due to two factors: the

ability to perform a wide range of in-line chemical pretreatment and sample handling

options, as well as compatibility with a diverse range of commonly used analytical

detection techniques[2]; which include: photometry, chemiluminescence, atomic

absorption spectroscopy, fluorescence, and electrochemical methods[3].

Of the aforementioned methods, photometry is the most commonly applied[4], owing

to two factors: there is a wide range of selective, sensitive and rapid

spectrophotometric reactions, and the relative ease of construction and cost-

effectiveness of photometric detectors. While there may be other detection methods

available for an analyte that demonstrate greater sensitivity or selectivity, photometry

is often chosen in preference due to the simplicity and accessibility of photometric

instrumentation and chromogenic reagents.

Photometric detection involves measuring the transmittance (T) of an incident light

beam (Po) by comparing the incident light intensity to the emergent beam intensity (P)

subsequent to traversing a given optical pathlength (b). Absorbance (A) is related to

transmittance logarithmically, as shown in Equation 3.1:

T = P/Po

A = -log(T) (3.1)


101

The concentration (c) of the absorbing species is related linearly to the absorbance

through the absorptivity coefficient (ε)[5]. This relation is called the Beer-Lambert

law (Equation 3.2):

A = εbc (3.2)

The instrumentation required for photometric detection is therefore straightforward: a

transparent vessel with a fixed pathlength, a light source, and a light intensity

detector. Providing analytical conditions are conducive to a linear relationship

between absorbance and analyte concentration (monochromatic light, mitigation of

stray light, low analyte concentration), the handling of data produced by photometric

detection is uncomplicated.

With the development of solid-state light emitting diodes (LEDs) and miniaturised

silicon photodiodes that are comparatively inexpensive and robust, photometric

detectors can be constructed that are ideally suited to the rigorous conditions

commonly encountered during field measurement. The economical nature of these

devices, from a standpoint of power consumption, size and weight, makes them

particularly attractive in the construction of portable instrumentation.

3.1.1 Flow cell design

Commercially available photometric flow-through cells typically feature a z- or u-

shaped flow configuration, and most commonly have an optical pathlength of 10

mm[6, 7]. Figure 3.1 illustrates the basic features of the z-cell design.


102

Figure 3.1 A schematic showing the fundamental design of z-configuration photometric flow-through cell. A light emitting diode shines a beam of light through the liquid path, with transmission being measured by a photodiode.

Ellis et al[8] discussed the limitations often encountered when employing this flow

cell design in FIA systems. Stray gas or bubbles may accumulate at the inlet and

outlet of the flow cell, which are usually at less than 90 degrees (z-configuration) to

the optical path. Trapped bubbles cause anomalies in the optical path due to light

scattering, which can cause a noisy baseline signal in addition to large signal spikes

upon their detachment. If a flow analysis system is designed for automated and

unattended field measurements, random trapping and detachment of bubbles will lead

to the collection of large amounts of unusable data, which seriously undermines the

reliability of the instrument.

Z-configuration cells also often exhibit limited sensitivity due to pathlength

restrictions (typically 10 mm) and sometimes large hydrodynamic dispersion. While


103

the volume of the illuminated section of the z-cell may be less than 20µL, the dead

volume of the cell is often several times this value, leading to increased dispersion.

Increasingly the light pathlength will lead to implicitly larger volumes, which causes

an increase in dispersion and a subsequent reduction in sensitivity.

In addition, z-configuration cells are particularly prone to refractive index effects[7].

These occur when the refractive index of the sample and carrier liquids are different,

which under laminar flow conditions causes the formation of elongated parabolic

lenses at the leading and trailing interface between the two zones. Light passing

through these lenses may be dispersed or focused, causing an aberration in the light

signal that can give rise substantial “ghost” peaks[9] even in the absence of absorbing

analytes. This effect may cause extensive errors in quantification if disregarded[10]. A

number of methods have been reported for the reduction of, or compensation for,

refractive index peaks. These include: measurement of the peak in the central section

of a large injected zone thus avoiding the refractive index interface present at either

end of the zone[11], subtraction of the refractive index peak using dual wavelength

spectroscopy[12, 13], in-line salinity compensation[10], and introduction of the beam

transverse to the axis of flow which reduces the lensing effect at the zone

interface[14]. However, these methods all have their own shortcomings. A sizeable

sample injection may result in reduced sample throughput. Introduction of the light

beam transverse to the axis of flow has an inherently shorter pathlength for the same

dispersion than a beam that passes longitudinally, causing insensitivity. Use of dual

wavelength spectrometry requires the use of a charge coupled device, or diode array

detector, or at the least multi-wavelength measurement. In-line salinity compensation


104

involves the use of flow injection methods with a somewhat more complex flow

manifold and tedious matrix-matching, especially in estuarine waters.

3.1.2 Multi-reflective flow cells

The use of mirror-coated helical[15] or straight capillaries[16] that exhibit multi-

reflective behavior as flow cells has been investigated as a means of surmounting the

shortcomings encountered when using the conventional z-configuration design. Ellis

and coworkers[8] developed a multi-reflective flow cell, consisting of a length of

circular glass capillary externally coated with a silver reflective surface. Two

windows were etched into the silvered surface to allow the introduction and collection

of a light beam at an angle of 60 degrees to the axis of flow. A light beam entering

this capillary would therefore undergo multiple reflections from one externally

mirrored sidewall to the next until the exit aperture is reached, as shown in Figure 3.2.

Figure 3.2 A schematic representation of the coated multi-reflective capillary; showing the introduction of the light beam, multiple reflections and collection of the emergent beam from the exit aperture. Reproduced from Ellis et al[8].

This cell was found to have improved sensitivity to a z-configuration cell of

comparative length, caused by two factors: a longer effective optical pathlength, and a


105

reduction in the hydrodynamic dispersion, as the cell consists of only a small diameter

capillary with minimal dead volume. Ellis et al[8] also reported a significant reduction

in the refractive index effect. The multi-reflective cell reduces the refractive index

effect because the light beam is introduced and propagated through the cell

transversely to the flow axis, and hence does not experience the full lensing effect,

whereas for the z-configuration cell, the light beam experiences the full lens effect

because the optical path is longitudinal to the axis of flow. This design is also far less

liable to trap bubbles as the connections between the cell and the flow analysis system

are also longitudinal to the flow axis, and hence any bubbles will simply pass through

the cell without becoming trapped.

A limitation of the particular multi-reflective cell design described by Ellis et al[8] is

that the coated silver sidewall mirror absorbs a certain fraction of the light beam

intensity upon each reflection[17]. The amount of absorption by the silver coating

depends on the light wavelength and the coating thickness, as shown in Figure 3.3.

50

60

70

80

90

100

300 330 360 390 420 450 480

Wavelength (nm)

% R

efl

ec

tan

ce Bare Al

Ag 111 Å

Ag 157 Å

Ag 218 Å

Ag 200 Å

Bare Ag

Figure 3.3 The percentage reflectance of silver and aluminium coating in comparison to irradiance wavelength and coating thickness. Reproduced from Sebag et al[18].


106

In the visible region (390 - 700 nm) a silvered surface will typically absorb around 5 -

10 % of the incident light upon each reflection[17, 18]. While the small amount of

light lost may be acceptable when working in the visible spectrum, the reflectance of a

silver coating decreases significantly as the near ultraviolet range is approached[18].

This precludes photometric detection in the ultraviolet using a silver coated multi-

reflective cell. Utilising an aluminium coating may improve reflective performance in

the ultraviolet (Figure 3.3); however, as much as 20 % absorption may still occur

upon each reflection when in the 200-400 nm range[19], which for a multi-reflective

cell will cause significant light attenuation even after only a few reflections.

3.1.3 Total internal reflective cells

The phenomenon called total internal reflection involves the complete internal

reflection of a beam of light when it strikes the boundary between two materials of

sufficiently differing refractive indices. Total internal reflection only occurs if a ray of

light is passing from a material of a higher refractive index to one of a lower

refractive index, and it strikes the medium interface at an angle greater than the

critical angle (θc). The critical angle is measured with respect to the normal to the

interface, and can be calculated from Equation 3.3:

θc = arcsin(n1/n2) (3.3)

where n1 is the refractive index of the less dense medium, and n2 is the refractive

index of the densest medium. This effect is illustrated in Figure 3.4


107

Figure 3.4 A diagram representing total internal reflection. When the angle of incidence is less than the critical angle, the majority of the light is refracted and a small amount reflected. When the incident angle exceeds the critical angle, the light beam is completely internally reflected.

As the difference between refractive indices at the air-glass interface of the external

wall of a glass capillary can be large (n1 = 1.003, n2 = ~1.3 - 1.5), it is possible to use

total internal reflection to propagate a light beam through a glass capillary. One of the

advantages of using total internal reflection at the air-glass interface over an

externally coated reflective surface is that a ray of light that undergoes total internal

reflection experiences absorption only by the glass it traverses, whereas a ray of light

reflected on a coated surface will experience some absorption at that surface in

addition to absorption occurring within the glass wall.

A simple model can be used to determine how much light is recovered from a multi-

reflective cell using total internal reflection at the external air-glass interface in

comparison to an identical cell except with an externally silver coated surface. GE214

fused silica quartz glass is assumed to absorb 10 % of the light every 10mm (based on

data from the Momentive Quartz product description) at 589 nm, and the silver


108

coating will be assumed to absorb 5% of the light for each reflection at 589 nm[18].

Assuming an incident angle of 45 o and 1.0 mm capillary wall thickness, a beam will

traverse 2.8 mm per reflection through the quartz, thus causing 2.8 % light attenuation

per reflection. A cell with an external silver wall coating will experience an additional

5% attenuation per reflection, for a total of 7.8 %. Ten reflections will be assumed.

Table 3.1 Data indicating the relative transmittance of a silver coated and uncoated total internally reflective cell. Multi-reflective

cell type Light attenuation per

reflection (E) Number of

reflections (i) Transmittance

T = (1-E)i

Total internal reflection 0.028 10 0.75

Silver coated 0.078 10 0.44

The data from Table 3.1 indicates that an uncoated cell offers superior light recovery

(75 %) in comparison to one coated with silver (44 %) at 589 nm. The disadvantage of

the total internal reflection over a coated surface is that the angle of light introduction

is restricted due to the requirements imposed by the critical angle of reflection, and

because of this the effective optical pathlength of the total internal reflection cell is

necessarily shorter over a designated capillary length. However, due to the greater

reflectance it is possible to include more reflections over a greater length of tubing,

which is the optical basis of liquid core waveguides.

Liquid core waveguide behavior enables the manufacture of very long pathlength

photometric cells. These waveguides consist of a low refractive index, polymetric

tubing material (e.g. Teflon AF2400®, or fused silica capillary coated either

internally or externally with the same polymer) with a flowing liquid core of higher

refractive index[20]. Total internal reflection is therefore achieved within the liquid


109

core at the polymer-liquid interface, with light propagating through the length of the

tubing until it reaches a collection point. These cells can be manufactured with optical

pathlengths in the order of metres[21], and provide considerably enhanced sensitivity

in comparison to traditional short pathlength z-type flow cells[22].

However, because the refractive indices of the coating, or polymetric material, and

water are often not that dissimilar; for example a Teflon AF2400 tube (n = 1.29) with

a water core (n = 1.33)[20], the critical angle is often very close to the axis of flow

(minimum θc = 76 o in the aforementioned case). Because the light beam propagates at

an angle almost parallel to the axis of flow, liquid core waveguides are very prone to

refractive index effects, which limit their applications in many flow injection analysis

techniques where multi-wavelength photometry is not possible.

The critical angle to achieve total internal reflection at the air-quartz interface of

GE214 Fused silica quartz capillary (n = 1.458, Momentive product description) can

be calculated to be 43.45 o (θc = arcsin[1.003/1.458]). In order to avoid the light

totally internally reflecting at the capillary-liquid interface, the angle must be

restrained to 66.09 o (θc = arcsin[1.333/1.458]). Thus, if the angle of incidence is

43.45 o < θ < 66.09 o, light will be reflected successively from each air-quartz

boundary and propagate through the cell, passing through the liquid core upon each

reflection[23], as seen in the optical simulation of Figure 3.5


110

Figure 3.5 An optical simulation of light undergoing total internal reflection within a circular quartz capillary, with an introductory angle of 51 o to the axis of flow. The larger cylindrical structures are areas at the ends of the capillary with the same refractive index, which are employed in the simulation as the means of introducing and collecting the reflected light beam. The cell dimensions are 0.555 mm wall thickness and 0.84 mm i.d.; the refractive indices of the cell material, air, and internal liquid are 1.458, 1.003, and 1.333 respectively.

As the incident angle of a quartz capillary is more transverse to the flow than in the

case of a liquid core waveguide, a cell of this design should exhibit a similar reduction

in refractive index effects to those reported by Ellis et al[8]. In order to introduce and

detect the internally reflected light beam, intentional light leakage points must be

constructed at reflection nodes (Figure 3.5). This may be easily achieved using a fibre

optic glued to the quartz surface using optical cement of a similar refractive index to

the quartz material of the capillary. These nodes may be predicted using simple

geometric modeling; these calculations are discussed in Section 3.2 of this chapter.

A total internally reflecting cell constructed from a quartz capillary has some of the

advantages of both liquid core waveguides (efficient light transmission, flexible

choice of irradiance wavelength) and coated multi-reflective capillary cells (reduced


111

refractive index effects, low dispersion, immunity to bubble entrapment) in addition to

the possibility of an extended pathlength in comparison to z-configuration flow cells.


112


• To design and construct a total internal reflective flow-cell suitable for use in a

flow injection system

• To determine the optical and hydrodynamic characteristics of the developed

total internal reflective cell

• To evaluate the relative tolerance of the total internal reflective cell to the

refractive index effect in comparison to a coated multi-reflective cell and a z-

configuration cell

• To compare the analytical performance (sensitivity, reproducibility, limit of

detection) of the total internal reflective cell, a coated multi-reflective cell and

a z-configuration cell using the same flow injection manifold for the

measurement of reactive phosphorus via the molybdenum blue method.


113

3.2 Experimental

3.2.1 Design and construction of flow cells

Total internal reflective cell

The total internal reflective cell was constructed from a 60 mm length of circular GE

214 fused silica quartz tubing (Momentive performance materials, Albany, New York,

USA) 1.95 mm o.d. and 0.84 mm i.d. mounted on a machined aluminium baseplate to

ensure the stability of the optics, as shown in Figure 3.6.

Figure 3.6 The total internal reflective cell capillary mounted on a metal stand. The terminals of the optical fibres are situated 18 mm apart, where they introduce and collect light internally reflected within the capillary.

Including the connecting tubing, the total volume of the cell was 71 µL. Two lengths

of quartz optical fibre of 1 mm internal diameter (P1000-2-UV/Vis, Ocean Optics,

Dunedin, Florida, USA) were mounted at 53 o to the normal of the capillary surface,

18 mm apart along the length of the tubing to introduce light and collect emergent

light. The fibres were cemented in place using N3 Norland UV curing optical

adhesive (Norland Products, Cranbury, New Jersey, USA). A red light emitting diode

diode (Serial#1513SRCE, λmax = 660 nm; 2800 mCd at 20 mA, Kingbright

Corporation, City of Industry, CA, USA) was used as the light source, and the


114

absorbance was measured using a USB-ISS-UV/Vis CCD detector (Ocean Optics,

Dunedin, FL, USA).

Tubular capillary multi-reflective and z-configuration cells

A 31 mm length of tubular borosilicate capillary (1.30 mm o.d. 0.80 mm i.d., total

volume 49 µL including connecting tubing) was externally coated with a silvered

surface via chemical deposition, as per the method described by Howard[24]. Two 1

mm diameter apertures were etched into the silver coating 13 mm apart to allow light

input and collection from the cell[8]. Ocean Optics 1 mm internal diameter quartz

fibres clamped at 30 degrees to the normal of the tubing surface were used to

introduce and collect light from a red light emitting diode, with the absorbance

measured using a USB-ISS-UV/Vis CCD detector (as in 3.2.1)

A 10 mm pathlength optical glass z-configuration cell (Starna Limited, Harnault,

Essex, UK, Model 75.15) with an internal diameter of 1.5 mm and volume of 18 µL

was used. The total volume of the cell including internal channels and connecting

tubing was 377 µL. The cell was mounted in a purpose built cell holder with optical

fibres positioned to the normal of the cell windows, using the same light source and

detector as in 3.2.1.


115

3.2.2 Reagents

Bromothymol blue dye

A stock solution of 1 gL-1 bromothymol blue was made by dissolved 0.1000 g of

Bromothymol blue sodium salt in 5 ml of ethanol, which was then diluted to 100 mL

using ultra pure water. This stock was then diluted using 0.01 M disodium tetraborate

decahydrate to make standards in the 1 – 10 mgL-1 range.

Reactive phosphorus standards

A stock solution of 100 mgL-1 phosphorus as orthophosphate was prepared by

dissolving 0.4394 g of potassium hydrogen phosphate in 1 L of ultra pure water. This

solution was refrigerated below 4 oC. An intermediate stock solution of 1 mgPL-1 was

prepared daily and used to make working standards of 10 – 100 µgPL-1.

Molybdenum blue chromogenic reagents

The acidic molybdate reagent was made by sonicating 5.00 g of ammonium

molybdate in ca. 250 mL ultrapure water until dissolved. 17.5 mL of concentrated

sulfuric acid was then added, and the solution made up to 500 mL with ultrapure

water. The acidic tin chloride reducing reagent was made by sonicating 0.10 g tin(II)

chloride and 1.00 g hydrazine sulfate in ca. 250 mL ultrapure water until dissolved.

14.0 mL of concentrated sulfuric acid was then added, and the solution made up to

500 mL with ultrapure water. Both solutions were stored for no longer than a week

and were sonicated for 10 minutes before use to outgas them.


116

Marine water for refractive index effects study

Low nutrient sea water, used to test the refractive index effect and for preparation of

phosphate standards, was collected from Port Phillip Bay at Mornington, SE

Australia, filtered using a 0.22 µm membrane (Acrodisc®, PALL Biosciences, Ann

Arbor, MI 48103, USA), and refrigerated at below 4 ◦C pending use. The filterable

reactive phosphorus content of this water was determined to be 21 µgPL-1.

3.2.3 Flow Injection Apparatus

An automated flow injection analysis instrument was used for the determination of

dissolved reactive phosphorus (Figure 3.7), and for the comparison of the three

different flow cells (Figure 3.8). The flow cells, including their interconnecting

tubing, were interchanged between comparisons. Two peristaltic pumps provided

liquid propulsion (Ismatec CA5E, Glattburg, Switzerland), an electrically driven valve

was used to make the sample injection (Model 5020, Rheodyne, Rohnert Park,

California, USA). System automation and data acquisition was handled using a

LabView® program on a personal computer, interfaced to the flow injection system

via a USB-1608FS Measurement Computing™ A-D DAQ board.


117

Figure 3.7 Flow injection apparatus for phosphorus used to evaluate the performance of the three cells. Flow rates; C = 1.6 mLmin-1, R1 = 1.1 mLmin-1, R2 = 0.7 mLmin-1. Figure 3.8 Flow injection apparatus for the detection of bromothymol blue used to evaluate the performance of the three cells. Flow rates; C = 2.1 mLmin-1.

3.2.4 Estimated pathlength of the, capillary multi-reflective and total internal reflective cell

The optical pathlength of the light entering a reflective flow-through cell can be

estimated by application of the laws of refraction and trigonometric functions [8, 16,

23] (Figure 3.9), providing the dimensions of the cell and the refractive indices of the

Cell

S

C

R1

R2

300 mm x 0.5 mm

600 mm x 0.5 mm

Injector

600 µL

Peristaltic pump

Or/Gre

Or/Wht

Or/Yel

Carrier = Ultra pure water R1 = Acidic Molybdate reagent R2 = Acidic Tin(II) Chloride reagent

Cell

S

C

Injector 250 µL

Peristaltic pump

Bl/Bl

Carrier = Ultra pure water


118

cell material and liquid contained therein are known to a reasonable degree of

accuracy.

Figure 3.9 A representation of light introduction and a single reflection in an externally coated capillary cell, reproduced from Ellis et al[8]. A total internal reflective cell behaves identically, save for the reflection occurring at an air-glass interface in place of a coated surface.

A ray passing from one medium to another will undergo refraction according to

Equation 3.4:

θ2 = arcsin[(n1sinθ1)/n2] (3.4)

The axial distance that the beam is displaced along the wall is then given by Equation

3.5:

l1 = d1 tan(θ2) = l3

l2 = d2 tan(θ4) (3.5)

and the total axial displacement for one reflection is given by Equation 3.6:


119

L = l1 + l2 + l3 (3.6)

In the silver coated cells, the reflection occurs at the outer glass wall, and similarly in

the total internal reflective cell the reflection occurs at the air-glass interface; thus

only the distance traveled through the liquid core contributes to the estimated optical

pathlength. This is calculated per reflection in Equation 3.7:

p = d2/cosθ4 (3.7)

The total optical pathlength (P) will be given by the ratio of the pathlength per

reflection (p) to the total axial displacement per reflection (L), multiplied by the total

distance between the inlet and outlet apertures (D), shown in Equation 3.8:

P = pD/L (3.8)

These calculations were performed for both the total internal reflective cell and the

coated capillary multi- reflective cell, and the results are shown in Table 3.2.


120

Table 3.2 Physical properties and optical parameters of the coated capillary multi-reflective and the total internal reflective cells. Parameters that are bolded are the input variables.

Physical cell parameter Symbol Coated multi-reflective

Total internal reflective

Refractive index of launch medium (589nm) n1 1.003 1.570

Refractive index of cell material (589nm)

n2 1.517 1.458

Refractive index of liquid (589nm)

n3 1.333 1.333

Cell wall thickness (mm) d1 0.25 0.555 Cell internal diameter (mm) d2 0.80 0.84

Angle of incidence (o) θ1 30 53 Reflection angle 2 (o) θ2 19.31 59.28 Reflection angle 3 (o) θ3 19.31 59.28 Reflection angle 4 (o) θ4 22.15 70.52 Reflection angle 5 (o) θ5 22.15 70.52 Reflection angle 6 (o) θ6 19.31 59.28 Ray length 1 (mm) l1 0.09 0.93 Ray length 2 (mm) l2 0.33 2.37 Ray length 3 (mm) l3 0.09 0.93

Total ray length (mm) L 0.50 4.24 Aperture Distance (mm) D 13 18

Number of reflections N 26.0 4.24 Optical length per reflection

(mm) p 0.86 2.42

Estimated pathlength through liquid (mm) P 22.42 10.69


121

3.3 Results and Discussion

3.3.1 Sensitivity of multi-reflective cells and accuracy of the estimated pathlength

When using photometric detection, there are three major parameters that affect the

sensitivity of a flow analysis system: the extent to which the chromogenic reaction

proceeds toward equilibrium, the optical pathlength of the flow-through cell and the

hydrodynamic dispersion of the operating manifold, which includes the intrinsic

dispersion of the flow-through detector. With respect to flow-cell design, the optical

pathlength and dispersion of the cell are the two parameters of interest. An increase in

optical pathlength and a decrease in dispersion would reasonably be expected to yield

an increase in sensitivity, and vice versa.

Experiments conducted by Ellis et al[8] showed that the coated capillary multi-

reflective cell was approximately 2.5 times more sensitive than a z-configuration cell

of 1.5 mm diameter aperture. Measurements also showed that sample passing through

the circular multi-reflective cell underwent approximately half the dispersion of that

measured in the z-configuration cell. Accordingly, Ellis et al[8] concluded that the

sensitivity improvement was due to both an increase in pathlength and a significant

decrease in dispersion. However, no other tests were performed in order to determine

to the extent to which each parameter affected the sensitivity increase, or to determine

the accuracy of the ray-tracing method employed to estimate the optical pathlength.

Standards of bromothymol blue in the range 1 – 10 mgL-1 were used to assess the

sensitivity and dispersion of the z-configuration cell, the circular multi-reflective cell


122

and the total internal reflective cell. The calibration regression equation, cell volume

and dispersion obtained in these measurements are listed in Table 3.3. The cell

volume represented in the table does include connecting tubing.

Table 3.3 Comparison of the sensitivity (calibration gradient) and dispersion of the Z, capillary multi-reflective and total internal reflective cells. Error values in the regression equation are calculated using least squares linear regression technique.

Cell Type Total Cell Volume

Calibration slope (mol-1L) Dispersion

Z-cell 377 µL 14030 ± 120 r2 = 0.9997 (n = 5)

1.63 ± 0.01 (n = 3)

MRC 49 µL 30540 ± 250 r2 = 0.9997 (n = 5)

1.47 ± 0.01 (n = 3)

TIR 71 µL 13940 ± 320 r2 = 0.9995 (n = 4)

1.47 ± 0.01 (n = 3)

The data in Table 3.3 indicate that the coated multi-reflective cell is the most sensitive

of the cells tested. As would be expected, dispersion increases as the total cell volume

increases. Similar to the value reported by Ellis et al[8], these experiments indicated

that the capillary multi-reflective cell was approximately 2.2 times more sensitive

than the comparative z-configuration cell.

Even though the total internal reflective cell exhibits multi-reflective behavior over a

capillary length of 18 mm, data in Tables 3.2 and 3.3 show that that the optical

pathlength and sensitivity of the cell are similar to that of the z-configuration cell. The

reason for this apparent contradiction is that for each reflection in the total internal

reflective cell, most of the path traversed by a ray is through the quartz capillary walls

rather than the liquid core. The capillary used for the coated cell had half the wall

thickness of the total internal reflective capillary (Table 3.2). Thus, the pathlength

through the liquid for a given length of capillary is dictated by the minimum critical


123

entry angle required to achieve total internal reflection and the ratio of the capillary

wall thickness to the internal diameter of the capillary. This is illustrated in Figure

3.10.

0.0

0.5

1.0

1.5

25 30 35 40 45 50 55 60

Entry angle (degrees)

Op

tical p

ath

len

gth

: C

ap

illa

ry len

gth Total internal reflective cell

Coated multi-reflective cell

Figure 3.10 A plot of the ratio of estimated optical pathlength to capillary length as a function of the light beam entry angle taken with respect to the normal of the flow axis. The total internal reflective cell and coated multi-reflective cell are normalised to the same physical dimensions (Table 3.1).

In Figure 3.10 the ratio of estimated optical pathlength to capillary length as a

function of the light beam entry angle indicates that even for the most acute entry

angles, the estimated optical pathlength for the total internal reflective cell is always

less than the length of the capillary. This is in marked contrast to the multi-reflective

cell, where the use of a more acute entry angle will result in an improvement to the

optical pathlength because of the increased number of reflections realised (Table 3.1).

Figure 3.10 also suggests that entry angles of greater than 53 o would result in an

improvement to the optical pathlength; however, in practice an angle of 53 o produced

the highest optical transmission through the cell, and therefore offers a greater signal


124

to noise ratio and improved limits of detection. Given that the light emitting diode

source is non-collimated, at entry angles greater than 53 o there are presumably

internal reflections within the capillary at the liquid-quartz interface that results in

reduced transmission as the upper critical angle bound of 66.09 o is approached. In

addition, as the angle of incidence is increased, the light beam propagates through the

liquid near parallel to the axis of flow (87 o for an entry angle of 66 o) and most likely

will exhibit similar refractive index effects as z-cells.

While measuring the gradient of a bromothymol blue calibration can yield a useful

comparison of the overall sensitivity of the different flow-through cells, this does not

reveal what measure of the sensitivity is due to a decrease in dispersion or an increase

in optical pathlength. According to the Beer-Lambert law (A = εbc), the gradient of

the calibration regression for each cell should be directly proportional to the

pathlength (b) and the dispersion, a pseudo-measurement of concentration (c), given

that the absorptivity of bromothymol blue (ε) remains constant between cells.

Given that the dispersion can be measured with a high degree of confidence, the

calibration regression equation gradient corrected for both the estimated optical

pathlength and the dispersion will provide an indication of the accuracy of the ray-

tracing measurement used to calculate the estimated optical pathlength of the two

reflective cells.


125

Table 3.4 Absorptivity values corrected for dispersion and pathlength and normalised the absorptivity coefficient (εmeas=22700 M-1cm-1) of bromothymol blue in 0.01 M borax, determined by batch method. The error in the estimated pathlength was estimated by adjusting the refractive indices of the liquid, launch material and cell material by ± 0.01 and the angle of incidence by ± 0.10 degrees. The error in the dispersion was calculated from the standard deviation of the triplicate measurements.

Cell Type

Calibration slope (S)

M-1

Estimated pathlength

(l) cm

Apparent molar abs. (έ = S/l) M-1cm-1

Dispersion (D)

n = 3

Molar abs. corrected

for D (εcorr= έD )

Normalised (εcorr : εmeas)

Z-cell 14030 ± 120 1.00 ± 0.00

14030 ± 120

1.63 ± 0.01

22856 ± 250

1.01 ± 0.01

MRC 30540 ± 250 2.24 ± 0.04

13622 ± 280

1.47 ± 0.01

19981 ± 420

0.88 ± 0.02

TIR 13940 ± 320 1.06 ± 0.01

13040 ± 346

1.47 ± 0.01

19169 ± 515

0.84 ± 0.02

In Table 3.4, the slope corrected for dispersion and estimated pathlength is normalized

against an experimentally determined absorptivity for the bromothymol blue dye in

borax used for the experiments. The normalised value for the z-configuration cell is

indicative of accuracy of these experiments as it has a definite pathlength of 1.00 cm,

and the corrected absorptivity value for the z-configuration cell normalised to the

known bromothymol blue absorptivity (1.01 ± 0.01) indicates the measurements are

highly accurate. The normalised values indicate that the ray-tracing method of

calculating the estimated pathlength of the reflective cells is therefore reasonably

accurate, with deviations of 16 (± 2) % for total internal reflective cell and 12 (± 2) %

for the coated multi-reflective cell.

There are three easily identifiable sources of error in the optical pathlength

estimations for the two reflective cells. The ray-tracing method assumes that the light

beam enters and exits the cell in a collinear fashion. A light emitting diode source is


126

non-collimated, and thus light will be introduced into the cell at a range of different

angles, as the optical fibre will emit light at the same angle at which it accepts from

the source. This can introduce a series of multiple light paths of differing lengths into

the cell. Multiple-path effects may also be exaggerated by the reflective nature of the

cells. Secondly, the ray-tracing method also assumes that reflection occurs only in one

plane. This assumption is reasonable if square capillaries are used; however, the use

of tubular capillaries causes rotation of the light rays about the longitudinal axis as

they propagate through the cell because of reflection from the curved walls. These

effects can be seen in the optical simulation in Figure 3.4. The refractive indices used

to calculate the estimated pathlength are specified for a single wavelength (589 nm);

however, an LED with a maximum emission wavelength of 660 nm was used in these

experiments, which gives rise to a degree of uncertainty in the refractive index values

used to estimate the optical pathlength.

3.3.2 Evaluation of the analytical performance of the z-configuration and reflective

cells using the photometric determination of reactive phosphorus

An operating FIA manifold was assembled for the detection of reactive phosphorus

using molybdate chemistry, as specified in 3.2.3, in order to further assess the

analytical performance of the total internal reflective cell in comparison with the z-

configuration cell and the coated multi-reflective cell. The manifold operating

conditions remained the same for each cell, save that the cells themselves were

exchanged between calibration experiments.


127

Table 3.5 Analytical performance of the three cells for the determination of reactive phosphorus. The reproducibility is measured using 100 μgPL-1

orthophosphate standard.

Cell Type Range (µgPL-1)

Linearity (R2)

Mathematical LoD*

Precision (% RSD)

(n=5) Z-Cell 10 - 100 0.9993 4.9 µgPL-1 0.80 MRC 10 - 100 0.9995 3.8 µgPL-1 0.27 TIR 10 - 100 0.9999 2.0 µgPL-1 0.66

*Limit of detection as determined by linear regression method used by Miller and Miller[25].

The measurements displayed in Table 3.5 indicate that the total internal reflection

flow-through cell has an analytical performance that is on-par with the other two cells.

An excellent detection limit of 2.0 µgPL-1 and %RSD of 0.66 (100 µgPL-1, n = 5)

show a significant improvement over the z-configuration cell and the coated multi-

reflective cell in terms of lower limit of detection. The reduced sensitivity of the total

internal reflective cell in comparison to the coated multi-reflective cell is somewhat

offset by its improved signal to noise ratio, which ultimately results in an improved

limit of detection.

3.3.3 Comparison of refractive index effects on the total internal reflective, coated

multi-reflective and z-cells

Refractive index effects occur when two zones of differing refractive index pass

through a flow-through cell, creating a parabolic lens that causes aberrations in the

light path, which leads to anomalous signals. A cell that exhibits a significant

reduction in these deleterious effects will consequently increase the accuracy of

photometric measurements in estuarine or marine waters. In order to determine the


128

extent of the refractive index effect on each of the cells, a 250 µL injection of nutrient

depleted sea water (n = 1.3394, S = 36.2) in an ultra pure water carrier stream (n =

1.3330) was interrogated spectrophotometrically to determine the size of the refractive

index peak produced.

-150

-100

-50

0

50

0 20 40 60 80

Time (seconds)

Dete

cto

r re

sp

on

se (

arb

itra

ry u

nit

s)

TIR Cell

Z-cell

MR Cell

Figure 3.11 The refractive index effect on the z-configuration cell, the circular coated multi-reflective cell (MR) and the circular total internal reflective cell (TIR), as determined by the injection of nutrient depleted sea water into an ultra pure water carrier.

Figure 3.11 indicates that the refractive index effect experienced by the two reflective

cells is relative small in comparison to that experienced by the z-configuration cell. Of

the reflective cells, the circular silver coated cell exhibited the most tolerance to the

refractive index effect, although there was still a noticeable effect. The marginally

greater refractive index effect exhibited by the total internal reflective cell is due to

the reflected light path which is more closely aligned to the longitudinal axis of flow

(70.5 o from the normal) which increases the lensing effect in comparison to the

coated multi-reflective cell (22.2 o from the normal) (Table 3.1). Predictably, the z-


129

cell which is configured such that the light beam axis is parallel to the axis of flow,

exhibits the greatest refractive index effect.

Standards in the range 10 - 100 µgPL-1 as orthophosphate in nutrient depleted marine

water were used to compare the extent of the peak distortion due to refractive index

effects in the total internal reflection and z-configuration flow-cells. Identical flow

injection manifolds were used, except that the flow cells were exchanged between

experiments.

-1000

-500

0

500

1000

1500

0 25 50 75 100

Time(s)

Dete

cto

r re

sp

on

se (

arb

itra

ry u

nit

s)

0

10

20

50

100

Figure 3.12 Flow injection peaks for orthophosphate in nutrient depleted marine water (concentrations 0 – 100 µgPL-1 as labeled on the chart) using the molybdenum blue method (660 nm) for the z-configuration cell.

As can be seen in Figure 3.12, there is a very pronounced refractive index effect

occurring in the peaks collected using the z-configuration cell. In estuarine or marine

waters, the z-configuration cell would produce erroneous measurements, particularly

at the lower concentration range (< 50 µgPL-1). The shape of the blank peak is

particularly distorted.


130

-200

0

200

400

600

800

0 25 50 75 100

Time (s)

De

tec

tor

res

po

ns

e (

arb

itra

ry u

nit

s)

0

20

50

100

10

Figure 3.13 Flow injections peaks for orthophosphate in nutrient depleted marine water (concentrations 0 – 100 µgPL-1 as labeled on the chart) using the molybdenum blue method (660 nm) for the total internal reflective cell.

However, there are no obvious refractive index effects observable in the peaks

recorded using the total internal reflective cell (Figure 3.13). A trendline plotted from

the peak heights indicates a high level of linearity (r2 = 0.9976). This data indicates

that spectrophotometric flow injection analysis methods can be performed on samples

of differing refractive indices without detriment to the accuracy using the total

internal reflective cell. The large blank peak recorded is due to phosphorus

contamination in the acidic molybdate and tin(II) chloride colorimetric reagents, as

well as some phosphorus present in the collected marine water and is not the result of

refractive index effects, which is quite evident upon comparison to the blank peak

shape in Figure 3.12.


131

3.4 Conclusion

A total internal reflective flow-through cell has been constructed for use in flow

injection analysis. The optical and hydrodynamic characteristics have been assessed

using bromothymol blue dye studies and ray-tracing techniques. The practical

advantages of this cell have been evaluated using the photometric determination of

reactive phosphorus by the molybdenum blue method. The aforementioned

characteristics of this cell were compared with a coated multi-reflective cell and a

conventional z-configuration cell. The total internal reflective cell has similar

sensitivity to the z-configuration cell, whereas the coated multi-reflective cell

exhibited greater sensitivity than either of these cells. For the determination of

reactive phosphorus, the FIA system equipped with a total internal reflective cell

achieved a superior detection limit compared with that using the coated multi-

reflective cell and the z-configuration cell, because of more efficient light

transmission and an accompanying higher signal to noise ratio. In comparison to the

z-configuration cell, the total internal reflective cell shows a markedly reduced

refractive index effect and an immunity to bubble entrapment. There is potential to

improve the analytical performance of the total internal reflective cell by use of a

capillary with a larger internal diameter to wall thickness ratio, or by increasing the

capillary length.

The ray-tracing method employed to determine the estimated optical pathlength for

the reflective cells[8, 16, 23] has also been shown to be reasonably accurate, with a

slight tendency to overestimate the optical pathlength (12 % for the multi-reflective

cell and 16 % for the total internal reflective cell). Possible explanations for this


132

discrepancy include the existence of multipath behavior, rotation of the light rays

around the cells longitudinal axis, and the use of a non-collimated source.

The total internal reflective cell is also more versatile than the coated multi-reflective

cell as it is not restricted to specific operational wavelength bands by absorbance or

scatter effects caused by reflective metal coatings. The total internal reflective cell

combines the reduced refractive index effect and immunity to bubble entrapment of

the multi-reflective cell with the higher signal to noise ratio and wider spectral range

of a conventional z-configuration cell. The versatility of the total internal reflective

cell has the potential to offer photometric detection within the ultra-violet spectral

range subject to the availability of suitable choices of light source, capillary material

and optical coupling media.


133

3.5 References 1. Ruzicka, J., and Hansen, E.L. (1988). Flow injection analysis, 2nd Edition

(John Wiley & Sons; USA).

2. McKelvie, I.D. (1999). Flow injection analysis. Analytical Testing


3. Hansen, E.H., and Wang, J. (2005). The three generations of flow injection

analysis. Analytical Letters 37, 345-359.

4. Christian, G.D. (2005). Optical methods comprise the majority of FIA

research and applications. TrAC, Trends in Analytical Chemistry 24, 560 -

563.

5. Skoog, D.A., and Leary, J.J. (1992). Principles of instrumental analysis, 4th

Edition (Floria: Saunders college).

6. Valcárcel, M., and Luque de Castro, M.D. (1987). Flow injection analysis—

principles and applications (Chichester: Ellis Horwood Limited).

7. Frenzel, W., and McKelvie, I.D. (2008). Photometric detection. In Advances

in flow injection analysis and related techniques, I.D. McKelvie, ed.

(Amsterdam: Elsevier), p. 311.

8. Ellis, P., Lyddy-Meaney, A.J., Worsfold, P.J., and McKelvie, I.D. (2003).



9. Dias, A.C.B., P. Borges, E.P., A.G. Zagatto, E.A.G., and Worsfold, P.J.

(2006). A critical examination of the components of the schlieren effect in

flow analysis. Talanta 68, 1076-1082.


134

10. McKelvie, I.D., Peat, D.M.W., Matthews, G.P., and Worsfold, P.J. (1997).

Elimination of the schlieren effect in the determination of reactive phosphorus

in estuarine waters by flow-injection analysis Analytica Chimica Acta 351,

265-271.

11. Yamane, T., and Saito, M. (1992). Simple approach for the elimination of

blank effects in flow-injection analysis of samples containing trace analyte and

an excess of another solute. Talanta 39, 215-219.

12. Liu, H., and Dasgupta, P.K. (1994). Dual-wavelength photometry with light

emitting diodes. Compensation of refractive index and turbidity effects in

flow-injection analysis. Analytica Chimica Acta 289, 347-353.

13. Zagatto, E.A.G., Arruda, M.A.Z., Jacintho, A.O., and Mattos, I.L. (1990).

Compensation of the schlieren effect in flow-injection analysis by using dual-

wavelength spectrophotometry. Analytica Chimica Acta 234, 153-160.

14. Jambunathan, S., Dasgupta, P.K., Wolcott, D.K., Marshall, G.D., and Olson,

D.C. (1999). Optical fiber coupled light emitting diode based absorbance

detector with a reflective flow cell. Talanta 50, 481-490.

15. Dasgupta, P.K. (1984). Multipath cells for extending dynamic range of optical

absorbance measurements. Analytical Chemistry 56, 1401-1403.

16. Wang, T., Aiken, J.H., Huie, C.W., and Hartwick, R.A. (1991). Nanoliter-

scale multireflection cell for absorption detection in capillary electrophoresis.


17. Bennett, J.M., and Ashley, E.J. (1965). Infrared reflectance and emittance of

silver and gold evaporated in ultrahigh vacuum. Applied Optics 4, 221–224.


135

18. Sebag, J., Krabbendam, V.L., Poczulp, G., Neill, D., Vucina, T., and Boccas,

M. (2006). LSST reflective coating studies (A.-E. Eli, A. Joseph and L.

Dietrich, eds.), vol. 6273. pp. 62730X, SPIE.

19. Hass, G., Heaney, J.B., Herzig, H., Osantowski, J.F., and Triolo, J.J. (1975).

Reflectance and durability of Ag mirrors coated with thin layers of Al2O3 plus

reactively deposited silicon oxide. Applied Optics 14, 2639–2644.

20. Byrne, R.H., and Kaltenbacher, E. (2001). Use of liquid core waveguides for

long pathlength absorbance spectroscopy: Principles and practice. Limnology

and Oceanography 46, 740-742.

21. Takiguchi, H., Tsubata, A., Miyata, M., Odake, T., Hotta, H., Umemura, T.,

and Tsunoda, K. (2006). Liquid core waveguide spectrophotometry for the

sensitive determination of nitrite in river water samples. Analytical Sciences,

1017 – 1019.

22. Wang, Z.A., Cai, W., Wang, Y., and Upchurch, B.L. (2004). A long

pathlength liquid-core waveguide sensor for real-time pCO2 measurements at

sea. Marine Chemistry 84, 73 – 84.

23. Tsunoda, K., Nomura, A., Yamada, J., and Nishi, S. (1989). The possibility of

signal enhancement in liquid absorption spectrometry with a long capillary

cell utilizing successive total reflection at the outer cell surface. Applied

Spectroscopy 43, 49-55.

24. Howard, N.E. (1969). Handbook for telescope making (London: Faber and

Faber Limited).

25. Miller, J.C., and Miller, J.N. (1993). Statistics for analytical chemistry

(Prentice Hall).

136

Chapter 4 – Ultra-violet spectrophotometric

flow analysis methods for the determination of

nitrate and total nitrogen in freshwaters

Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters

137

4.1 Introduction

4.1.1 Nitrogen in natural waters

The stoichiometry of the Redfield equation (C:N:P = 106:16:1) suggests that nitrogen,

along with carbon and phosphorus, is an essential nutrient for primary production[1].

Increases in nitrogen concentrations in natural waters may accelerate harmful algal

blooms commonly associated with eutrophication[2]. In some cases, nitrogenous

species may be growth limiting[3], particularly in waters where denitrifying bacterial

activity is prevalent[4]. Nitrogen is often thought to be the limiting nutrient for

photosynthetic growth in marine waters, and hence is a commonly monitored nutrient

in natural waters; especially anthropogenic inputs of nitrogen, such as treated effluent

discharges[5].

There are several operational categories that are used to classify fractions of the total

nitrogen concentrations in waters. Dissolved inorganic nitrogen (DIN) are those

species that will pass through 0.45µm filter[4] and includes nitrate, nitrite and

ammonia, which are considered to be the most bioavailable forms of aquatic

nitrogen[4]. Since nitrate and nitrite are often determined simultaneously, these two

species are often measured concurrently and reported as NOX[6]. Total nitrogen is the

measurement of all nitrogen within a body of water, i.e. nitrogen in colloidal and

particulate matter, within organisms and dissolved in waters. Measurement of

dissolved inorganic nitrogen provides an indication of the amount of immediately bio-


138

available nitrogen, whereas quantification of total nitrogen yields an estimate of the

total potentially bioavailable nitrogen.

Total nitrogen concentrations can often be less than 10 µgNL-1 in coastal and open

marine waters and pristine freshwaters[7], but can exceed 1000 µgNL-1 in systems

affected by anthropogenic contamination[8]. Sewage effluents have been known to

contain > 35 mgNL-1 and typically contribute of up 60 % of the total nitrogen

concentration of the receiving waterway[8]. ANZECC guidelines (1992)[9], which

were formulated to protect water quality, recommend that the TN concentration of

rivers and streams should fall in the range 100 - 750 µgNL-1, whereas for marine and

coastal waters NOX concentrations should be 1 - 60 µgNL-1.

4.1.2 Techniques for measuring dissolved inorganic nitrogen species in natural

waters

Methods for the direct detection of nitrate include ultra-violet spectrophotometry[10-

12], nitrate specific ion-selective electrodes[13-15] and ion chromatography[4]. Ultra-

violet spectrophotometry offers the benefits of using instrumentation common to

many analytical laboratories, as well avoiding the use of chromogenic reagents and

thus eliminating potentially noxious or expensive reactants[4]. As nitrate absorbs

strongly in the 200 - 230 nm region[10], quantification by ultra-violet

spectrophotometry is prone to interferences from chloride, and to a lesser extent

bromide, which are found in high concentrations in marine-estuarine waters[12], as

well as from natural organic matter in freshwaters[16]. However, measurements at

reference wavelengths can be used to correct for the aforementioned interferences[12,


139

16] as well as the use of polynomial correction functions[17]. Second derivative ultra-

violet spectrometry has also been successfully used to eliminate interference from

organic matter without the requirement of multiple wavelength correction[3, 18-20].

Ion selective electrodes specific to nitrate have been used effectively in the

determination of nitrate in fresh and wastewaters[13], although anions commonly

found in natural waters (chloride, bicarbonate, bromide, iodide, nitrite) cause

significant interference[13, 21], which severely limits the potential applications of

these electrodes. In addition, the lower detection limit of these potentiometric methods

is reported to be 140 µgNL-1, which is too insensitive for application in many natural

waters[21]. Ion chromatographic methods find limited application for analysis of

marine waters due to the high ionic strength of these samples and interference from

high concentrations of chloride[4]. Residual oxidant from the digestion of total

nitrogen may also cause damage to ion chromatography columns.

The most commonly used method for the quantification of nitrite involves

diazotization in the presence of sulfanilamide, followed by reaction with a coupling

agent such as N-(1-napthyl)ethylenediamine dihydrochloride, to form an intensely

coloured azo dye that can be measured using spectrophotometric detection[16, 21,

22]. This reaction is referred to as the Griess Assay. The absorption maximum for the

azo-dye occurs between 500 and 600 nm, depending on the coupling agent[22, 23].

Nitrate may also be determined by this method following the complete reduction of

the nitrate to nitrite, in which case the sum of nitrite and nitrate is determined, which

is referred to as NOX[23].


140

Multiple methods have been reported for the reduction of nitrate to nitrite in waters;

including reduction with solid copperised cadmium granules, wires or tubes[23-27],

alkaline aqueous hydrazine[28-32], photo-induced reduction using ultra-violet

light[33-39], and enzymatic reduction with nitrate reductase[40, 41].

Quantification of nitrate by reduction using a cadmium column coupled with the

Griess reaction was first reported in 1960[24], and since then has become the most

commonly used method for the measurement of NOX in waters, due primarily to its

sensitivity and relative freedom from sample matrix effects[25]. The reduction of

nitrate in the presence of cadmium occurs according to Equation 4.1[24]:

NO3-(aq) + H2O(l) + 2e- → NO2

-(aq) + 2OH-

(aq) Eo = 0.0100V (4.1)

The reduction of nitrate is accompanied by the oxidation of the cadmium, as shown in

Equation 4.2[24]:

Cd(s) → Cd2+(aq) + 2e- Eo = 0.403V (4.2)

The reaction occurs most efficiently under alkaline conditions, and quantitative yields

on a rapid time-scale require a pH greater than 8[23, 26]. The reducing ability of

cadmium may be increased by coating the column with a metal of lower reduction

potential. Cadmium columns treated with a copper sulfate solution, in order to

precipitate a porous layer of copper on the cadmium metal surface, have been reported

to increase the rate of the redox reaction up to 8 fold[24, 27].

While reduction of nitrate with cadmium columns coupled with spectrophotometric

detection through the Griess reaction is a selective, rapid and sensitive method, there


141

are some disadvantages. Cadmium metal is highly toxic, and prolonged exposure

represents significant risk to the operator, as well as posing the difficulty of disposal

of noxious Cd2+ waste. Phosphate, in concentrations greater than 3 mgPL-1, will

inhibit nitrate reduction due to binding to the Cd-Cu particles[42]. Cadmium columns

also have a limited lifetime, with the reduction efficiency of the column decreasing

over time due to the passivation of the cadmium surface by the formation of cadmium

hydroxide and cadmium carbonate[24]. McKelvie et al[2] also found that residual

oxidant from the digestion of organic nitrogen compounds using peroxodisulfate

significantly reduced the column lifetime, which limits the application of these

columns for the determination of total nitrogen.

Hydrazine has been proposed as an alternative reductant to copperised cadmium[28,

39]. The reaction requires higher temperatures[29] (up to 70oC) and alkaline

conditions[30] in order to provide quantitative reduction. The reduction reaction is

shown in Equation 4.3[28]:

2NO3-(aq) + N2H4(aq) ↔ 2NO2

-(aq) + 2H2O(l) + N2(g) (4.3)

Nitrite produced by this reaction is also commonly measured using the Griess

reaction[28, 30, 39]. However, there are several disadvantages to hydrazine reduction

that limit its potential applications to natural waters using flow analysis techniques;

namely, nitrate reduction is inhibited by precipitation of magnesium ions from

marine-estuarine and brackish waters under the alkaline reactions conditions[30, 31],

a by-product of hydrazine reduction is the evolution of dinitrogen gas[28] that can

impede spectrophotometric detection through bubble entrapment, and reaction times


142

of up to 24 hours can be required in order to gain complete conversion of nitrate to

nitrite[32].

Nitrate can also undergo ultra-violet photo-induced reduction[33-39]. At pH 8 the

reduction proceeds directly to nitrite, as shown in Equation 4.4[35]:

NO3-(aq) + hν → NO2

-(aq) + ½O2(g) (4.4)

In-line photo-reduction of nitrate has been employed in flow analysis techniques, with

sample throughputs ranging from 6 to 25 samples per hour for a variety of sample

matrices[35, 38]. However, there is also the potential for photo-oxidation of other

nitrogen containing compounds to nitrite in the presence of dissolved oxygen[39].

Oxygen is produced within the photo-reactor due to photolysis with water, which may

cause re-oxidation of nitrite to nitrate[35, 38].

Enzymatic reduction of nitrate using nitrate reductase has also been reported as a

potential alternative to techniques involving cadmium[40, 41]. Nitrate reductase is an

oxidoreductase that catalyses the redox reaction of nitrate to nitrite in the presence of

nicotinamide adenine dinucleotide (NADH), as shown in Equation 4.5[43]:

NADH(aq) + NO3-(aq) + H+

(aq) → NAD+(aq) + NO2

-(aq) + H2O(l) (4.5)

The enzymatic reaction operates most efficiently under mildly acidic conditions i.e

pH 5.5 – 7[40, 41]. Under optimum conditions, the reactions kinetics are slow and

even automated flow systems may be limited to 10 samples per hour[43]. In addition,

enzymes are often expensive and require stringent storage conditions.


143

Nitrate and nitrite may also be reduced to ammonium, by a number of processes

including; reduction using zinc[44-46], Devarda’s alloy[6, 47, 48], and titanous

chloride[49-51]. The ammonia formed can then be quantified by several methods;

spectrophotometrically via the Berthelot reaction[52], gas sensing ammonia

potentiometric electrodes[49], spectrophotometrically after gas diffusion of ammonia

into a pH sensitive colorimetric indicator[53], and by direct detection through gas-

phase ultra-violet spectrophotometry[51].

Reduction of nitrate and nitrite using zinc and Devarda’s alloy are problematic for

flow analysis applications. Devarda’s alloy increasingly absorbs magnesium and

calcium hydroxide with time[47, 54], and consequently the alloy must be renewed

regularly or sampling time must be increased in order to maintain complete reduction.

Copious amounts of hydrogen gas are also produced as a result of reduction using

Devarda’s alloy[48]. Copperised granular zinc was found to have a longer lifetime

than Devarda’s alloy; but, the alkaline conditions required cause precipitation of

magnesium and calcium ions present in natural waters, causing clogging of the zinc

column[45]. Titanous chloride, while being effective at reducing nitrate and nitrite to

ammonia, is highly toxic and produces harmful fumes[50], in addition to undergoing

rapid deterioration upon exposure to air[55]. Titanous chloride is not favoured as a

reductant because it is a risk to operator safety and its inflexible storage requirements

are incompatible with field use.


144

4.1.3 Techniques for digestion of total nitrogen

Due to their recalcitrant nature, many dissolved and particulate nitrogenous species

are difficult to measure directly. In order to quantify total nitrogen, all nitrogen

containing compounds must first be converted to a more easily detectable form, such

as nitrate or ammonia. This process is called mineralisation or digestion and may

involve dissolution, oxidation or hydrolysis, or any combination thereof depending on

the nature of the sample. For flow analysis, mineralisation is usually achieved by an

automated in-line digestion step. Assuming complete mineralisation of all nitrogenous

compounds to nitrate or ammonia, a measurement of the produced mineral species

performed on the digested sample can be used to quantify the total nitrogen

concentration.

The Kjeldahl method, first reported 1883 by Johan Kjeldahl[56], is historically the

most common method of determining organic nitrogen, and is still in wide use

today[21, 57-59]. The Kjeldahl digestion involves heating the sample in the presence

of highly concentrated sulfuric acid, causing the mineralisation of organic nitrogenous

species to ammonium hydrogen sulfate, according to Equation 4.6[60]:

Organic N(aq) + H2SO4(aq) → CO2(g) + H2O(l) + NH4HSO4(aq) (4.6)

Following complete digestion, the sample is treated with an excess of sodium

hydroxide that liberates the generated ammonia for quantification[61]. While the

Kjeldahl method effectively mineralises many nitrogenous species of biological origin

such as proteins, peptides and amino acids, it is far less effective at converting nitro (a

large component of the total nitrogen pool) and cyano compounds to ammonia[62].


145

Therefore, the Kjeldahl digestion cannot be used to provide a measurement of total

nitrogen concentration[62]. The use of harsh acidic conditions and highly toxic

reagents also present significant operator and environmental risk. In addition,

Kjeldahl digestion times can be quite lengthy[2].

In response to the shortcomings of the Kjeldahl determination of nitrogenous

compounds, several alternative digestion procedures have been proposed for the

determination of total nitrogen. Digestion methods based on ultra-violet irradiation

were first reported in 1966 by Armstrong et al[63] and autoclaving of samples with

alkaline peroxodisulfate was reported by Koroleff[64] in 1969. Procedures to quantify

total nitrogen discussed in the literature include; photo-oxidation in the presence of

small volumes of hydrogen peroxide[63, 65-68], high temperature combustion in the

presence of oxygen gas[7, 69, 70], thermal alkaline peroxodisulfate digestion[3, 71-

76] including microwave induced digestion[77], and combined photo-oxidation

alkaline peroxodisulfate techniques[2, 5, 78-80].

The procedure introduced by Koroleff[64] used alkaline peroxodisulfate as an

oxidant. The aqueous peroxodisulfate decomposes upon exposure to heat (100 - 120

oC) in an autoclave, according to Equation 4.7[73]:

S2O82-

(aq) + H2O(l) → 2HSO4- (aq) + ½O2(g) (4.7)

It is reported that the oxidation of organic nitrogen compounds to nitrate occurs due to

oxidative reaction with the oxygen that is liberated upon the thermal decomposition of

peroxodisulfate[73, 75]. Measurement of total nitrogen can be obtained by

quantifying the nitrate present in the digested sample[3, 71-76]. Nitrate is the sole


146

product of this digestion using alkaline conditions; however, acidic digestion

conditions cause the formation of multiple unidentified nitrogenous compounds[81].

The thermal alkaline peroxodisulfate method is considered superior to the Kjeldahl

procedure because it measures organic nitrogen and free ammonia as well as nitrate

and nitrite[72]. While the alkaline peroxodisulfate method offers advantages over the

Kjeldahl digestion, lengthy digestion times (in the order of 15 - 90 minutes for batch

methods) are reported[71-73]. The copious amounts of oxygen bubbles produced may

also prove problematic for spectrophotometric detection in flow analysis methods if

the thermal digestion is performed in-line.

As an alternative to wet chemical methods, several high temperature combustion

procedures have been reported[7, 69, 70]. Sea water samples are heated in a furnace in

the presence of pure oxygen, with temperatures from 670[69] to 1110 oC[7] being

used. The combustion products include nitric oxide which can be measured via

chemiluminescence[7], and nitrogen dioxide which is detected using the Griess

reaction[69]. Although this approach provides excellent conversion of nitrogenous

compounds, the severe reaction conditions required precludes its application for

shipboard total nitrogen analysis[69]. In addition, the high temperatures and vapour

phase reactions would be difficult to achieve using an in-line thermal reactor for flow

analysis.

The photo-oxidative procedure reported by Armstrong et al[63] involved irradiating

marine samples using a mercury arc lamp (λmax=254nm) in the presence of a small

amount of hydrogen peroxide[63], thus converting nitrogenous compounds to nitrate,


147

which can then be measured to quantify total nitrogen. This procedure was

subsequently further applied in marine, estuarine[65, 66] and fresh waters[66, 68].

Upon ultra-violet irradiation, hydrogen peroxide will decompose, producing hydroxyl

radicals that are highly reactive with organic compounds[82]. This decomposition can

be seen in Equation 4.8:

H2O2(aq) + hν → 2OH•(aq) (4.8)

The hydroxyl radicals oxidise nitrogenous species to nitrate. Armstrong et al[63]

found that nitrogen containing compounds were oxidised to nitrate over a period of 2 -

3 hours, and further irradiation lead to the reduction of nitrate to nitrite.

This method has been predominantly applied using batch analysis[63, 65, 66, 68],

where the samples are irradiated in silica tubes. The use of hydrogen peroxide as a

hydroxyl radical source in flow analysis methods employing in-line digestion has

been limited because copious amounts of oxygen bubbles are produced upon

hydrogen peroxide decomposition, which can often impede spectrophotometric

detection of nitrate.

Peroxodisulfate, while a strong oxidant, reacts slowly with many organic species[83].

Similar to hydrogen peroxide, upon exposure to ultra-violet radiation, a

peroxodisulfate medium will produce hydroxyl and sulfate radicals, which are strong

oxidising agents. Equation 4.9 shows the radical generation process[83]:

S2O82-

(aq) + hν → 2SO4-•

(aq)


148

2SO4-•

(aq) + H2O(l) � HSO4-(aq) + OH•

(aq) (4.9)

The hydroxyl and sulfate radicals then may react with organic compounds, further

decompose peroxodisulfate (Equation 4.10), or undergo reactions with other radicals

(Equation 4.11):

S2O82-


-(aq) + SO4

-•(aq) + 1/2O2(g) (4.10)

SO4-•


-(aq) + 1/2O2(g) (4.11)

Both the sulfate and hydroxyl radicals are responsible for the destruction of organic

compounds. Either radical may dominate this process depending on the digestion pH,

with the hydroxyl radical being produced principally under alkaline conditions and

sulfate radical production occurring primarily under acidic conditions[83]. Thus, a

peroxodisulfate ultra-violet method can operate successfully under either acidic or

alkaline conditions. The advantages of using an alkaline medium include; the partial

suppression of carbon dioxide generated from the oxidation of organic compounds,

complete conversion of nitrogenous compounds to nitrate only[81], whereas acidic

oxidising conditions have been reported to reduce the conversion efficiency of

ammonium[5], a species that can be a substantial component of the total nitrogen

pool.

Alkaline peroxodisulfate photo-oxidation using automated by flow injection analysis

offers significantly faster digestion than Kjeldahl, batch thermal alkaline

peroxodisulfate methods, and hydrogen peroxide based photo-oxidative techniques;

with McKelvie et al[2] reporting a frequency of 25 samples per hour, Roig et al[5]

measuring 20 samples per hour, and Cerda et al[78] reporting a sample throughput of

12 per hour. Recoveries exceeding 90 % of refractory nitrogenous compounds, such


149

as urea, glycine, nicotinic acid, and aspartic acid have been achieved using alkaline

peroxodisulfate photo-oxidative methods[2, 5, 78].

The automated flow injection method developed by McKelvie et al[2] utilised a

copperised cadmium column to reduce the nitrate generated by the photo-oxidative

digestion to nitrite, followed by spectrophotometric detection using the Griess

reaction. However, residual peroxodisulfate from the in-line digestion of total

nitrogen was found to cause a white precipitate on the surface of the cadmium

granules (presumably cadmium hydroxide) that was accompanied by a gradual

decrease in the reduction efficiency of the copperised cadmium column. McKelvie

and coworkers[2] attempted to eliminate the residual oxidant through reaction with

sodium metabisulphate, but this caused a large blank signal and only delayed column

degradation rather than preventing it completely. In addition to shortening the lifetime

of cadmium reduction columns, the presence of residual oxidant could also be

problematic for any metal-based nitrate reduction technique, such as the zinc and

Devarda’s alloy procedures mentioned previously in this chapter. Hence, direct

methods of determining nitrate may be preferable to those involving a reduction step

of nitrate, whether to nitrite or ammonia.

4.1.4 Direct measurement of nitrate in the presence of residual peroxodisulfate

Two reported methods for direct measurement of nitrate are ultra-violet spectroscopy

and nitrate-specific ion selective electrodes. Nitrate specific electrodes offer poor

selectivity and sensitivity[13, 21], and therefore their application to natural waters is

limited. Furthermore, the presence of residual oxidant is likely to be deleterious to the


150

electrode membrane. However, direct photometric detection of nitrate using ultra-

violet spectroscopy has the potential to be a simple and reagent-free method for the

determination of digested nitrogenous species. The measurement of nitrate ions is

usually achieved using wavelengths between 200 - 230 nm[10], with corrections for

interferences found in natural waters, such as chloride or organic matter, taken at

reference wavelengths closer to the visible i.e. 275 nm[12].

However, residual peroxodisulfate from the oxidation of nitrogenous species to nitrate

also absorbs strongly in the 200 - 230 nm region, which potentially limits this

approach to the detection of high concentrations of nitrogen (> 10 mgNL-1). There

have been three proposed methods to overcome the spectral interference caused by the

residual oxidant, i.e.; second derivative ultra-violet spectroscopy[3], spectral

deconvolution[5], and single-wavelength measurement using the residual

peroxodisulfate signal as a blank[79].

As the absorbance from the non-interacting contributing species is additive, the

concentration of individual components at any given wavelength can be isolated using

the Beer-Lambert law. Equation 4.12 demonstrates this process for theoretical non-

interacting absorbing species Z and X at two separate wavelengths λ’ and λ”[84];

A’ = ε’xbcx + ε’zbcz at λ’

A” = ε”xbcx + ε”zbcz at λ” (4.12)

where the absorptivity coefficients ε’ and ε” are evaluated from the slopes of the

individual calibration curves of the relevant species at λ’ and λ” respectively, and the


151

known cell pathlength is b. Spectral deconvolution methods enable determination of

nitrate concentration against significant spectral overlap from residual oxidant and

other interfering species by considering the ultra-violet absorbance spectra as a linear

combination of non-interacting overlapping spectra (as demonstrated in Equation

4.12) from individual species[5]. The sample spectrum (SW) is the sum of several

reference spectra, as indicated in Equation 4.13[5]:

SW = ∑ai REFi ± r (4.13)

where ai is the contribution coefficient of ith reference spectrum REFi, and r is the

quadratic error. While Roig et al[5] reported excellent accuracy using this

deconvolution method, no natural samples with a total nitrogen concentration below

10 mgNL-1 were examined; and as such, the performance of this method in detection

low concentrations of nitrogen (<1 mgNL-1) in the presence of high residual oxidant

concentrations is not known.

Hinkamp and Schwedt [79] used single wavelength detection, at 226 nm, to determine

nitrate after photo-oxidative digestion using peroxodisulfate. A large blank signal

occurred due to strong absorbance by the residual oxidant, and consequently the

detection limit of the method was limited to 0.75 mgNL-1, making this technique

generally too insensitive for the determination of total nitrogen in natural waters.

Thomas et al[80] suggested that a longer irradiation time, of around 15 minutes, could

be used to photolyse the residual peroxodisulfate, and thus reduce the size of the blank

signal. However, an increase in irradiation time would significantly reduce sample

throughput, as well as generating a larger number of oxygen gas bubbles.


152

Second derivative ultra-violet spectroscopy has been applied in batch methods for the

determination of nitrate following autoclave digestion using alkaline

peroxodisulfate[3, 85]. A large proportion of the peroxodisulfate oxidant undergoes

thermal decomposition during autoclaving, and hence the second derivative

spectroscopy method requires no correction for residual oxidant or sulfate produced

by thermal decomposition (Equation 4.7), and provides excellent accuracy[3, 85].

This data analysis approach also allows quantification of nitrate in the presence of

organic matter and high concentrations of phosphate[85]. However, there is an

inherent increase in noise upon derivitisation of absorbance spectra[86], causing the

signal to noise ratio to decrease with higher order derivatisation.

The research discussed in this chapter investigates the photo-oxidation of nitrogenous

compounds in natural waters to nitrate using a ultra-violet photo-reactor with alkaline

peroxodisulfate. Direct ultra-violet spectrophotometric determination of nitrate was

evaluated as an alternative to the more commonly utilised cadmium reduction of

nitrate to nitrite followed by the Griess reaction. Investigations of multi-wavelength

background correction and second derivative spectroscopy as a means of overcoming

the spectral overlap caused by residual peroxodisulfate are described.


153


The design, construction and evaluation of flow analysis instrumentation for the

measurement of total nitrogen and nitrate in natural waters are described in this

chapter according to the following objectives:

• To utilise in-line photo-oxidation as a means of digesting nitrogenous

compounds to nitrate using flow analysis method

• To investigate direct ultra-violet spectrophotometric approaches to the

quantification of nitrate in natural waters as well as nitrate generated during

digestion of total nitrogen

• To determine if increased photo-reactor irradiation time significantly reduces

the residual peroxodisulfate signal, and to evaluate the effectiveness of a

multi-wavelength background method and second derivative ultra-violet

spectroscopy for the elimination of the peroxodisulfate blank signal

• To design and construct a single-reflection capillary flow-through cell to

minimise entrapment of oxygen gas bubbles generated from the decomposition

of peroxodisulfate


154

4.2 Experimental

4.2.1 Reagents

Alkaline peroxodisulfate digestion agent

Potassium peroxodisulfate (0.50 g) and disodium tetraborate (0.50 g) were dissolved

in ultrapure water up to 100 mL. The pH of this solution was 9.1. The influence of

peroxodisulfate concentration was investigated using 1.25, 2.5 and 10.0 gL-1 solutions

all of which were buffered to pH 9.1. These solutions had a lifetime of approximately

7 days at room temperature, after which there was a noticeable decrease in their

oxidising ability.

Nitrate standards

In a 500 mL volumetric flask, 0.3035 g of sodium nitrate was dissolved in 500 mL

ultrapure water to make a 100 mgNL-1 nitrate stock solution. This solution was

refrigerated at 4 oC and diluted as appropriate.

Nitrite standards

In a 500 mL volumetric flask, 0.2465 g of sodium nitrite was dissolved in 500 mL

ultrapure water to make a 100mgNL-1 nitrite stock solution. This solution was

refrigerated at 4 oC and diluted as appropriate.


155

Model nitrogen compounds

Stock solutions of each of the model nitrogen compounds were prepared to a

concentration of 100 mgNL-1, by dissolving the solid in 500 mL of ultrapure water

using a volumetric flask, followed by refrigeration at 4 oC. These included 0.1911 g

ammonium chloride, 0.6647 g ethylenediaminetetraacetic acid sodium salt, 0.2681 g

glycine, 0.4397 g nicotinic acid, and 0.1073 g urea.

Artificial seawater

1000 mL of artificial seawater was prepared as per the method described by Kester et

al[87].

Collection of water samples

All samples were collected unfiltered from various locations around storm water

drainage in Clifton Hill, Victoria, Australia over four different time periods in

November-December 2009. The samples were stored frozen until measured.

4.2.2 Instrumentation

Sampler and digestion module

A sampler and digestion module was used to handle all sample treatment operations;

including digestion, debubbling and filtration. This module was identical to that

described in Section 2.2.2, except that the electric heating unit was bypassed, with

ultra-violet treated sample being pumped directly into the hollow fibre filter. Labeled

photographs of the sampler-digestion module (Figure 2.2-3) is shown in Section 2.2.2.

Figure 4.1 is a schematic diagram of the sampling and digestion module. Automation


156

of the digestion module functions was achieved using a USB-1608FS Measurement

Computing™ A-D DAQ board, interfaced to a personal computer running a LabView

(v. 8.5) control and data acquisition program.

Figure 4.1 A schematic representing the digestion module and the single reflection continuous flow detection system. Volumes and flow-rates are listed. The single reflection cell consists of a rectangular capillary with a single aluminium coated surface, where light from a ultra-violet source undergoes a single reflection before entering a charge coupled device.

Single-reflection flow-through cell for ultra-violet spectrophotometric measurement

A 75 mm length of rectangular (external dimensions 6.6 x 5.5 mm, internal

dimensions 4.0 x 2.1 mm) GE 214 fused silica quartz tubing (Momentive performance

materials, Albany, New York, USA) was externally coated with aluminium by

vacuum deposition along the 5.5 mm wide external face. The coated length of tubing

UV Reactor

Digestion reagent, 2 mLmin-1

Sample in, 2 mLmin-1

Waste

Hollow-fibre filter, 300 µL

Debubbler

Waste, 0.3 mLmin-1

To flow-cell, 2.3 mLmin-1

2000 mm, 0.8 mm i.d., 1000 µL

Waste Peristaltic

Pump

Digestion Module

Detection Module

Al reflective cell

UV source

CCD


157

was fixed in a dove-tail slide that could be adjusted using a worm-drive to reach

maximum light transmission, as shown in Figure 4.2.

Figure 4.2 The single-reflection flow-through cell, featuring an external coated aluminium reflective surface. Light is introduced at 45 o and undergoes a single reflection from the aluminium surface before emerging from the capillary.

Two quartz optical fibres (P1000-2-UV/VIS, Ocean Optics Inc, Dunedin, FL, USA,

1000 µm diameter) were mounted at 45 o to the normal of the tubing surface 11 mm

apart. One of the optical fibres introduces light from a 30 W deuterium ultra-violet

source (J16T, Applied Biosystems, Carlsbad, CL, USA), while the other collected the

emergent light beam and guided it into a USB-ISS-UV/Vis CCD detector (Ocean

Optics, Dunedin, FL, USA). The light beam underwent a single reflection within the

cell from the opposing mirror coated wall. This cell has an approximate pathlength of

10.4 mm. The large dead volume of this cell does not reduce the sensitivity of this

method because there is a continuous flow of digested liquid through the cell rather

than injection of a sample zone that would undergo dispersion. If stop-flow is

employed, the volume of the photo-reactor (ca. 1000 µL) is large enough to produce

an undispersed irradiated zone in the cell which has a volume of ca. 630 µL. A


158

peristaltic pump (Ismatec CA5E, Glattbrugg, Switzerland) was used to pump liquid

from the debubbler on the sampler module through the flow cell at 2.3 mLmin-1. Data

acquisition was handled by OOIBase32 software (Ocean Optics, Dunedin, FL, USA)

via a USB interface with a personal computer.


159

4.3 Results and Discussion

The direct ultra-violet photometric measurement of nitrate is typically undertaken in

the 220 – 230 nm range. Second derivative data analysis methods have been used to

overcome any spectral interference from organic matter in natural waters and

phosphate commonly found in high concentration in waste waters[20]. In addition,

saline waters also contain interfering species (e.g. chloride, bromide)[12]. In order to

determine whether ultra-violet spectrophotometric detection of nitrate in marine

waters is viable, an evaluation of the extent of the spectral interference from chloride

is necessary.

4.3.1 Interference of chloride for ultra-violet measurement of nitrate

Chloride and bromide are reported to be significant spectral interfering species in the

determination of nitrate, both by direct ultra-violet detection[12] and second

derivative methods[20]. Of these anions, chloride is present in large concentrations in

marine waters and estuaries (typically 19300 mgClL-1 in marine waters) to prove a

significant interference. To evaluate this potential interference, the spectra of artificial

seawater at various stages of dilution (25, 50, and 75 %) were collected. The second

derivative spectrum of the 25 % artificial seawater solution was also calculated. These

spectra are compared with the absorbance and second derivative spectra of a 1

mgNL-1 nitrate standard in Figure 4.3.


160

0

0.25

0.5

0.75

1

1.25

190 210 230 250

Wavelength (nm)

Ab

so

rban

ce

-0.018

-0.012

-0.006

0

0.006

0.012

d2A

/d_

2

25 % Artificial Seawater 50 % Artificial Seawater

75 % Artificial Seawater Artificial Seawater

1 mgN/L Nitrate 2nd Derivative 25 % Artificial Seawater

2nd Derivative 1 mgN/L Nitrate

Figure 4.3 Spectra of various dilutions of artificial sea water and a 1 mgNL-1 as nitrate solution.

Even when diluted, the chloride present in the artificial seawater is a significant

spectral interference in the 200 – 230 nm range in which nitrate is commonly

determined. The second derivative spectrum in Figure 4.3 also indicates that chloride

has a similar effect for second derivative spectra. Given the obvious unsuitability of

this method for marine-estuarine waters, all future investigation was conduced in

fresh waters.

4.3.2 Measurement of nitrate in freshwaters using second derivative spectroscopy

The basis of the second derivative spectrophotometric technique for quantifying

nitrate is that the maximum rate of change that occurs for nitrate absorption at around

225 nm is unique to that species[20], and as such eliminates interference from any


161

organic matter, phosphate, metal species (such as iron and copper) found at

concentrations typically expected in natural waters[85]. A nitrate standard of 10.0

mgNL-1 is used to illustrate this procedure in Figure 4.4 below.

-3.5

0

3.5

215 225 235 245

Wavelength (nm)

Ab

so

rban

ce

-0.25

-0.125

0

0.125

0.25

De

riv

ati

ve

va

lue

Absorbance dA/d_ d_A/d__

Figure 4.4 The absorbance spectra of a 10.0 mgNL-1 as nitrate standard, with the first and second derivative also shown. The first and second derivative spectra are plotted on the secondary axis.

Figure 4.4 indicates a clear second derivative peak at around 227 nm. As derivatising

spectra is known to degrade the signal to noise ratio[86], quite heavy smoothing (a 10

point moving average) has been applied to the first and second derivative curves

shown in Figure 4.4.

In order to determine if the second derivative method is selective for nitrate or if it

measures both nitrite and nitrate (NOX), two 1 mgNL-1 standards for nitrate and nitrite

and a additional 1 mgNL-1 standard consisting of a 1:1 mixture nitrite and nitrate were

measured, and their second derivative spectra calculated (Figure 4.5).


162

-0.004

-0.002

0

0.002

210 220 230 240

Wavelength

d2A/d_2

1mgN/L Nitrate 1mgN/L Nitrite 0.5mgN/L Nitrate & 0.5mgN/L Nitrite

Figure 4.5 The second derivative spectra of 1 mgNL-1 nitrate and nitrite standards, along with a standard consisting of a 1:1 mixture of the two.

Figure 4.5 indicates that it is possible to measure NOX where the three second

derivative spectra intercept at approximately 235 nm. However, if the NOX

concentration is primarily nitrate, then a measurement in the 225 nm region offers

superior sensitivity. Considering that nitrite concentrations in freshwaters are typically

very low, being present mostly as an intermediate in the bacterial denitrification

process[32], a wavelength of 226 nm was chosen to monitor nitrate concentration.

To evaluate the analytical performance of the second derivative method, standards in

the range 0.0 - 2.0 mgNL-1 as nitrate were measured using continuous flow system

with a single-reflection flow cell (Figure 4.6).


163

-1500

0

1500

3000

4500

0

Second Derivative Peaks

d2A

/d_

2 (

x1

07)

Blank 0.25 mgN/L 0.50 mgN/L 1.00 mgN/L 2.00 mgN/L

Figure 4.6 Second derivative peaks for nitrate standards in the 0.0 - 2.0 mgNL-1 range.

The analytical figures of merit calculated from this calibration data are listed in Table

4.1 below.

Table 4.1 The analytical figures of merit for the second derivative nitrate method. The limit of detection is determined by the linear regression method described by Miller and Miller[88]. Sensitivity 2.13*10-4(d2A/dλ2)/mgNL-1

Precision (%RSD for 1 mgNL-1 nitrate) 0.4% (n = 10) Limit of Detection (99% conf. limit) 0.04 mgNL-1

Limit of Quantification (10σblank) 0.05 mgNL-1 Linearity (0.0 – 2.0 mgNL-1) R2 = 0.9995

As there is no chromogenic reaction involved in this determination, any variability

originates primarily through fluctuations in the deuterium light source and through an

increase in the noise that is inherent in data treatment such as derivatisation. While the

sensitivity may appear to be low, there is enough information in the second derivative

spectra to obtain gradation to three significant figures (Figure 4.6). While the 0.04


164

mgNL-1 lower detection limit is slightly inferior to that offered by cadmium

reduction-Griess methods (< 0.01 mgNL-1[24]), the second derivative method offers

the advantage of higher tolerance to phosphate, elimination of the need for toxic

reducing agents and colorimetric reagents. Additional experiments have shown that

the second derivative method produces a highly linear calibration for nitrate

concentrations less than 3 mgNL-1. Because there is no reduction or chromophoric

chemistry required for this method, the rate at which samples can be measured

depends on the speed of the peristaltic pump and the rate at which the software can

acquire the ultra-violet spectra.

The nitrate concentration of 20 natural freshwater samples was measured using the

second derivative spectroscopy technique. The only pretreatment undertaken was to

filter the samples using a 0.22 µm filter. The digestion module was bypassed and the

samples collected using the peristaltic pump connected directly to the flow cell

(Figure 4.1), where they were measured in triplicate in less than 5 seconds. The

salinity for each sample was determined using a salinity meter and used as an indirect

measure of the chloride concentration in order to determine the effect of chloride

interference on the second derivative determination of nitrate. The results obtained are

compared to those obtained using a cadmium reduction/Griess reaction method

(Figure 4.7).


165

y = (0.9538 ± 0.0203)x

+ (56.70 ± 20.20)

R2 = 0.9937

0

500

1000

1500

2000

0 500 1000 1500 2000

Nitrate (!gNL-1

) by comparative method

Nit

rate

(!

gN

L-1

) b

y s

eco

nd

deri

vati

ve

me

tho

d

Salinity < 0.1 0.2 > Salinity > 0.1

Figure 4.7 A comparison of nitrate concentration (µgNL-1) as determined by the second derivative method and a comparative method (cadmium reduction-Griess assay). Two series, all samples below and those above 0.1 salinity, are plotted to indicate the extent of the chloride interference. There are four points above 0.1 salinity, two that are overlapping. The pink line represents a 1:1 comparison.

Nitrate concentration determined by the second derivative method gives an excellent

correlation with nitrate determined by cadmium reduction-Griess method. A

Wilcoxon signed rank test (n = 16, ptwo-tail = 0.093) also indicates there is no overall

bias in the comparison. However, when sample salinity exceeds 0.1, the chloride

interference begins to cause significant error in the determination (Figure 4.7).

Accordingly, the second derivative method is limited to freshwaters with salinity

below 0.1 when measuring nitrate in this range (0 - 2 mgNL-1). Additional

experiments indicate that dilution of the sample along with a 2 mgNL-1 nitrate spike

can increase the chloride tolerance to salinity < 0.5.


166

4.3.3 Interference of residual peroxodisulfate in the ultra-violet measurement of

digested total nitrogen

It has been reported that residual peroxodisulfate absorbs strongly in the 200 – 230

nm region in which nitrate is commonly measured using ultra-violet

spectrophotometry[5, 79, 80]. In order to determine the extent of this interference, the

spectra of 0 - 2 mgNL-1 nitrate standards in both ultrapure water and 2.5 gL-1 alkaline

peroxodisulfate were obtained (Figure 4.8).

0.0

0.5

1.0

1.5

2.0

200 220 240 260

Wavelength (nm)

Ab

so

rbam

ce

UPW & P'sulfate 0.5 mgN/L & P'sulfate 1.0 mgN/L & P'sulfate

2.0 mgN/L & P'sulfate 0.5 mgN/L & UPW 1.0 mgN/L & UPW

2.0 mgN/L & UPW

Figure 4.8 Ultra-violet spectra of nitrate standards (0 - 2 mgNL-1) with ultrapure water (UPW) and 2.5 gL-1 alkaline peroxodisulfate (P’sulfate).

There is significant spectral overlap between nitrate and peroxodisulfate in the 200 –

230 nm region. The peroxodisulfate, which is present in far higher concentration than

nitrate, also causes significant suppression of the nitrate signal (Figure 4.8). While


167

there is a clear gradation of the nitrate signal with concentration in the presence of

peroxodisulfate, Figure 4.8 indicates that the sensitivity of this method would prove to

be quite poor. In addition, the standard deviation (n = 3) of the peroxodisulfate in

ultrapure water sample is equivalent to approximately 0.1 mgNL-1 of the nitrate signal

at 220 nm, which severely degrades the detection limit of this method.

Thomas et al[80] suggested that use of extended irradiation times, of up to 15

minutes, could decrease the concentration of residual peroxodisulfate, and thus reduce

the extent of the interference caused by the remaining oxidant. Figure 4.9 shows the

far ultra-violet absorbance spectra of a 2.5 gL-1 peroxodisulfate solution after ultra-

violet irradiation for various intervals of time.

0

0.5

1

1.5

2

200 220 240 260

Wavelength (nm)

Ab

so

rban

ce

0 min 3 min 6 min 10 min 15 min 30 min

Figure 4.9 A 2.5 gL-1 peroxodisulfate solution exposed to ultra-violet irradiation from a medium pressure ultra-violet lamp for 0 - 30 minute periods of time.

Figure 4.9 shows that an approximate 60 % reduction of the peroxodisulfate signal at

220nm occured after exposure to ultra-violet radiation for 10 minutes or longer.


168

Peroxodisulfate is known to decompose upon exposure to heat (Equation 4.7) and also

upon exposure to ultra-violet radiation (Equation 4.9). Figure 4.5 indicates that this

process reaches completion after approximately 10 minutes. As the ultra-violet lamp

and photo-reactor are housed in a stainless steel tube, the photo-reactor also heats up

to ca. 70 oC during periods of stopped flow, which may also assist thermal

decomposition of peroxodisulfate during the irradiation interval. The effectiveness of

the debubbling system employed in the sampler-digestion module ensures that oxygen

bubbles evolved from the decomposition of peroxodisulfate are successfully removed

from the stream, and do not impede spectrophotometric detection.

In order to determine whether the sensitivity of the method could be improved by

decomposing the residual peroxodisulfate, 0.0 - 2.0 mgNL-1 nitrate standards in 2.5

gL-1 peroxodisulfate were irradiated for 15 minutes. The ultra-violet spectra of the

irradiated standards are shown in Figure 4.10.

0.0

0.4

0.8

1.2

200 212.5 225 237.5 250

Wavelength (nm)

Ab

so

rban

ce

Blank 0.5 mgN/L 1.0 mgN/L 2.0 mgN/L

Figure 4.10 Ultra-violet spectra of nitrate standards (0.0 - 2.0 mgNL-1) with 2.5 gL-1 peroxodisulfate after irradiation for 15 minutes.


169

The increased gradation between the nitrate standards seen in Figure 4.10 shows that

reducing residual peroxodisulfate significantly increases the methods sensitivity, in

comparison with those measurements shown in Figure 4.8. This approach also

substantially reduces the impact of the peroxodisulfate blank signal on the lower limit

of detection, with one standard deviation (n = 3) being equivalent to 0.03mgNL-1 in

this case. The differences in sensitivity and limit of quantification are summarised in

Table 4.2.

Table 4.2 A summary of the differences in sensitivity and limit of quantification with increased ultra-violet irradiation time.

Method Type Sensitivity (at 220 nm)

Limit of Quantification (10σblank) (n = 3)

Continuous flow irradiation 0.058 A/mgNL-1 1.0 mgNL-1

15 mins stop-flow irradiation 0.103 A/mgNL-1 0.3 mgNL-1

The data in Table 4.2 indicates that a stop-flow approach dramatically improves the

methods sensitivity and limit of quantification. However, the limit of quantification

for the stop-flow method (0.3 mgNL-1) is still inadequate for natural waters. The cause

of the high limit of quantification is the large contribution that the residual

peroxodisulfate makes to the additive absorbance (Figure 4.10). Repeated

measurements of the peroxodisulfate blank signal indicate there is a degree of

imprecision (RSD = 1.1%, n = 10) in the blank signal, which while seeming relatively

small, is still significant in comparison to the contribution of nitrate to the total signal.

Therefore, fluctuations in the peroxodisulfate signal must be corrected for in order to

improve the lower detection limit of this method. Given that nitrate does not absorb

above 240 nm (Figure 4.8), it should be possible to correct for small fluctuations in

the blank signal using the absorbance of peroxodisulfate in the 245 nm region. If a


170

series of measurements of peroxodisulfate at slightly different concentrations and

different irradiation times are undertaken, a constant (K) between the absorbance in

the 220 nm region and the 245 nm region that corresponds to peroxodisulfate can be

obtained, as in Equation 4.14:

A(nitrate)220 nm = A220 nm – KA245 nm (4.14)

While this method was highly successful at subtracting the residual peroxodisulfate

signal for standards in ultrapure water, difficulties were encountered in real samples

where species other than peroxodisulfate also absorb in the 245 nm region (Figure

4.11). Even small contributions from unknown species at 245 nm caused significant

error upon subtraction of the background signal.

0

0.25

0.5

200 215 230 245 260

Wavelength (nm)

Ab

so

rban

ce

Blank 0.5 mgN/L 1.0 mgN/L 2.0 mgN/L Freshwater Sample

Figure 4.11 Ultra-violet spectra of nitrate standards (0 – 2 mgNL-1) and a freshwater sample in 2.5 gL-1 peroxodisulfate are irradiated for 15 minutes. The freshwater sample has interfering species absorbing in the 245 nm region that prevent the estimation of residual peroxodisulfate in this region.


171

Using a library of ten real samples, the absorbance at 275 nm, where peroxodisulfate

does not absorb, was used to further modify the constant K in Equation 4.14 in

addition to the measurement at 245 nm. However, this introduced further error into

the measurements. While the multi-wavelength background correction method

described here is similar to the successful spectral deconvolution procedure reported

by Roig et al[5], the deconvolution method requires a large difference between the

absorbing species (Equation 4.12). Roig et al[5] achieved this by measuring

wastewater in which the total nitrogen concentration was never less than 10 mgNL-1.

Multi-wavelength background correction is clearly more difficult for samples in the

0.0 - 2.0 mgNL-1 range.

Due to the aforementioned difficulties, second derivative spectroscopy was

investigated, which was previously been applied to the determination of nitrate in

fresh waters[3, 20, 85].

4.3.4 Measurement of total nitrogen using second derivative spectroscopy

As the second derivative spectroscopy method is capable of measuring nitrate in low

chloride freshwaters with a high degree of accuracy, and does not suffer from the

same interferences that the direct ultra-violet method does[20], further evaluation of

its potential application for the measurement of mineralised nitrogenous compounds

in the presence of residual peroxodisulfate was undertaken.

Initially, an investigation of the tolerance of the method to residual peroxodisulfate

was performed. A series of alkaline potassium peroxodisulfate solutions (0.63, 1.25,


172

2.5 and 5.0 gL-1) were irradiated for 0, 5, 10, 15 and 30 minutes in order to determine

at what point the peroxodisulfate decomposition was complete, and thus determine the

minimum irradiation time required to obtain the highest sensitivity and repeatability.

0

0.6

1.2

1.8

0 10 20 30

Irradiation Time (mins)

Ab

so

rban

ce

0.63 g/L 1.25 g/L 2.5 g/L 5.0 g/L

Figure 4.12 The effect of irradiation time on peroxodisulfate decomposition. Absorbance was measured at 220 nm. Error bar are ± 1 σn-1 for n = 3.

Figure 4.12 indicates that peroxodisulfate undergoes decomposition until an

irradiation time of approximately 10 minutes is reached. The second derivative

spectra are calculated from the 2.5 gL-1 peroxodisulfate data collected. The results are

shown in Figure 4.13


173

-0.01

-0.005

0

0.005

210 220 230 240

Wavelength (nm)

d2A

/d_

2

0 mins 5 mins 10 mins 15 mins 30 mins

Figure 4.13 Second derivative spectra of residual peroxodisulfate after different irradiation times. The chart indicates that after 10 minutes of irradiation, the blank signal in the 220 – 225 nm region where nitrate is measured remains uniform.

When only a short irradiation time is employed (< 10 mins), the residual

peroxodisulfate interferes significantly in the 220 – 225 nm region where the

derivative peak for nitrate is measured. However, as the irradiation time is increased,

this effect decreases until the blank signal becomes constant after 10 minutes. This

greatly improves the ease of ultra-violet measurement of nitrate, as no background

correction is required, provided that an irradiation time of at least ten minutes is used.

A comparison of the data in Table 4.2 shows clearly that the use of ultra-violet

radiation to decompose residual peroxodisulfate significantly increases the sensitivity

of the nitrate signal when direct ultra-violet measurement is employed. To determine

if this same effect was noted in the second derivative method, a 2.5 gL-1 alkaline

peroxodisulfate solution was irradiated for various time intervals (0, 5, 10, 15 and 30

minutes) both with ultrapure water and a 1.0 mgNL-1 nitrate standard. The sensitivity


174

suppression (i.e. the ratio of the blank corrected second derivative nitrate peak

obtained in peroxodisulfate to the second derivative nitrate peak measured in ultrapure

water) is determined from this data.

0

25

50

75

100

0 10 20 30

Irradiation time (mins)

% S

en

sit

ivit

y s

up

pre

ssio

n

Figure 4.14 The effect of irradiation time on the sensitivity of second derivative nitrate detection in the presence of 2.5 gL-1 peroxodisulfate. The chart indicates that irradiation time does not significantly increase sensitivity; however it does markedly improve repeatability i.e. signal to noise ratio. Error bars are ± 1 σn-1 for n = 3.

The sensitivity suppression mimics the pattern seen in Figure 4.12, decreasing as

residual peroxodisulfate decreases. The data in Figure 4.14 shows that an irradiation

time of 30 minutes only increases the relative sensitivity by 13%. This indicates that

irradiation time does not have a significant effect on the sensitivity of the second

derivative method. However, the standard deviation of the blank signal is much higher

for shorter irradiation times, which has an adverse influence on the detection limit of

this method.

The effect of peroxodisulfate concentration on the sensitivity of the second derivative

methods was also investigated. A series of alkaline potassium peroxodisulfate


175

solutions (0.63 - 5.0 gL-1) were irradiated for 15 minutes, either mixed with 1.0

mgNL-1 as nitrate or with ultrapure water. The percentage sensitivity suppression is

achieved by .comparing the blank corrected nitrate peak obtained in the presence of

peroxodisulfate to the nitrate peak measured with ultrapure water.

0

25

50

75

100

0 1.25 2.5 3.75 5

Peroxodisulfate (gL-1

)

% S

en

sit

ivit

y s

up

pre

ssio

n

Absorbance 2nd Derivative

Figure 4.15 The effect of peroxodisulfate concentration on the sensitivity of direct ultra-violet and second derivative nitrate detection in the presence of 5.0 gL-1

peroxodisulfate irradiated for 15 mins. The second derivative method is more sensitive in the presence of peroxodisulfate than the direct ultra-violet method. Error bars are ± 1 σn-1 for n = 3.

Figure 4.15 indicates that initial peroxodisulfate concentration is more important than

irradiation time in terms of maximising sensitivity. It is also worth noting that the

sensitivity suppression experienced by the second derivative method is far less than

the suppression experienced in non-derivatised absorbance spectrum.

In order to determine the digestion efficiency for conversion of various nitrogenous

compounds to nitrate, 5 model compounds (ammonium chloride,

ethylenediaminetetraacetic acid, glycine, nicotinic acid and urea) of concentration 1


176

mgNL-1 were measured after digestion and compared to a nitrate standard of

equivalent concentration using 0.63, 1.25 and 2.50 gL-1 alkaline potassium

peroxodisulfate digestion agents. A stop-flow irradiation time of 10 minutes was

chosen, according to the data in Figure 4.13.

0

25

50

75

100

Ammonia EDTA Gylcine Nicotinic Acid Urea

% c

on

versio

n

0.63 g/L Peroxodisulfate 1.25 g/L Peroxodisulfate 2.50 g/L Peroxodisulfate

Figure 4.16 The percentage conversion of various 1 mgNL-1 model compounds. With the exception of ammonium chloride, the percentage conversion of the model compounds improves significantly as the peroxodisulfate concentration increases. Error bars are ± 1 σn-1 for n = 3.

Figure 4.16 indicates that the best conversion was achieved using a 2.5 gL-1 alkaline

peroxodisulfate digestion agent. All model compounds were converted to nitrogen

with an efficiency exceeding 90 %, except for urea which gave 88 % (± 1.2 %)

conversion. Increasing the peroxodisulfate concentration may further improve the

conversion of urea, however a longer irradiation time would be required (Figure 4.12)

to reduce the residual peroxodisulfate signal causing a decrease in throughput.

Following the optimisation, operating conditions of 2.5 gL-1 peroxodisulfate oxidant

and an irradiation time of 10 minutes were adopted in order to maximise sensitivity,

precision, and to ensure high conversion of nitrogenous species to nitrate. Under the


177

aforementioned conditions, a calibration using nitrate standards in the range 0.0 - 2.0

mgNL-1 was used to determine the analytical figures of merit for the developed flow

analysis system (Table 4.3).

Table 4.3 The analytical figures of merit for the photo-oxidative total nitrogen method using second derivative detection. The limit of detection is determined by the linear regression method described by Miller and Miller[88]. Sensitivity 1.62*10-4 (d2A/dλ2)/mgNL-1

Precision (%RSD 1 mgNL-1 ammonia) 1.2% (n = 10) Throughput 5h-1, measured in triplicate Limit of Detection (99% conf. limit) 0.05 mgNL-1

Limit of Quantification (10σblank) 0.06 mgNL-1 Linearity (0.0 – 2.0 mgNL-1) R2 = 0.9989

Similarly to the second derivative method for nitrate in freshwaters, any imprecision

originates primarily from fluctuations in the deuterium light source and through an

increase in the noise that is inherent in data treatment such as derivatisation. While the

sample throughput is low compared to that reported by workers using a cadmium

column coupled with the Griess reaction (25 samples per hour[2]), this method

eliminates the toxic reduction column and problems associated with its reduced

lifetime in the presence of residual oxidant, and is thus simpler and more reliable

particularly over long-term periods. ANZECC guidelines (1992)[9] suggested that the

total nitrogen content of rivers and streams should be in the range of 0.10 - 0.75

mgNL-1, and as such a limit of detection of 0.05 mgNL-1 should be sufficient to

quantify total nitrogen content in all but the most pristine freshwaters.

In order to evaluate the accuracy of the developed method in natural waters, ten

freshwater samples were analysed for total nitrogen and the results were compared

with those determined by a autoclave alkaline peroxodisulfate method coupled with


178

cadmium reduction and the Griess reaction. The only sample pretreatment performed

was filtration using a 100 µm nylon mesh as to remove any particles large enough to

block the manifold tubing. The total nitrogen concentration of the samples was in the

range 0.60 - 2.60 mgNL-1 (Figure 4.17).

y = (0.9423 ± 0.0380)x

+ (0.0345 ± 0.0605)

R2 = 0.9872

0

1

2

3

0 1 2 3

Total nitrogen comparative method (mgNL-1

)

To

tal

nit

rog

en

se

co

nd

de

riv

ati

ve

me

tho

d (

mg

NL

-1)

Figure 4.17 A comparison of total nitrogen concentration as determined by the second derivative method and a comparative method (autoclave in alkaline peroxodisulfate followed by cadmium reduction-Griess assay). The pink line represents a 1:1 comparison. Error bars are ± 1 σn-1 for n = 3.

The comparison in Figure 4.17 indicates that there is strong agreement between the

two methods. A Wilcoxon signed rank test (n = 10, ptwo-tail = 0.160) also indicates

there is no overall bias in the comparison. While there is more scatter for second

derivative total nitrogen determination than is present for nitrate determination

(Figure 4.11), this is not unexpected as quantification of total nitrogen requires an

additional in-line digestion step. The strong agreement between the two methods also

confirms that conversion of nitrogenous compounds to nitrate proceeds to completion

under the experimental conditions.


179

4.4 Conclusion

The aim of the work detailed in this chapter was to develop methods for the

quantification of nitrate and total nitrogen using direct ultra-violet spectrophotometric

measurement. Direct photometric measurement was chosen over the commonly used

cadmium reduction technique due to the toxic materials and poor life-time of packed

copperised cadmium columns. Primary research problems reported when using ultra-

violet spectrophotometry to measure total nitrogen include spectrophotometric

interferences from residual peroxodisulfate oxidant and from naturally occurring

species such as chloride, bromide and organic matter.

A spectral second derivative method for determining nitrate in freshwaters was

evaluated. This method was found to be free from of interference from organic matter

commonly encountered during ultra-violet spectrophotometric measurements in

freshwaters. The method was simple; requiring only a ultra-violet transparent flow-

cell, a ultra-violet light source and a charge coupled device detector. As the method

involves no chromogenic agents, the precision is excellent (0.4 %RSD, n = 10, 1.0

mgNL-1 as nitrate). The lower detection limit of 0.04 mgNL-1 is suitable for most

natural waters. Comparative analysis undertaken on 20 freshwater samples indicated a

strong degree of agreement between the second derivative method and a cadmium

reduction method, except for those samples above salinity 0.1, as chloride is a

significant spectral interferent (Figures 4.3 and 4.7). Thus, this method is only

applicable to freshwaters of salinity less than 0.1 when measuring in the 0.0 - 2.0

mgNL-1 range.


180

Preliminary investigations into the measurement of total nitrogen involved developing

a multi-wavelength method to correct for residual peroxodisulfate, as the fluctuating

background could cause significant error in the nitrate determination if left

uncorrected. However, the wavelength region used to correct of the residual oxidant

(245 nm) was also shared by unknown absorbing species in natural waters, which

caused an error in the correction leading to largely skewed nitrogen values. A third

wavelength was chosen that was free of absorbance from peroxodisulfate and nitrate

(275 nm) to correct for the unaccounted for absorbance at 245 nm. While this did

improve the accuracy of the method in natural waters, there were still unacceptable

errors (up to ± 30 %) involved.

The second derivative method was investigated for the detection of nitrate generated

during the digestion of total nitrogen. It was found the second derivative method

significantly reduced both the peroxodisulfate blank signal and the imprecision

associated with it, provided a stop-flow irradiation time of at least ten minutes was

used. The second derivative nitrate peak was also found to undergo less sensitivity

suppression than the direct nitrate peak when in the presence of residual

peroxodisulfate. A concentration of 2.5 gL-1 alkaline potassium peroxodisulfate was

found to yield greater than 90 % conversion for 4 nitrogen model compounds

(ammonium chloride, EDTA, glycine and nicotinic acid) and 88 % for urea using an

irradiation time of 10 minutes. The lower detection limit 0.05 mgNL-1 is adequate for

most freshwaters. Comparative analysis undertaken on 10 freshwater samples

indicates strong agreement between the developed method and an autoclave based

comparative method (Figure 4.17).


181

While the second derivative total nitrogen method is limited to freshwaters and has a

reduced throughput (5 digestions per hour measured in triplicate) in comparison to

reported flow analysis based total nitrogen methods using a cadmium reduction

column and the Griess reaction (25 measurements per hour[2]), this procedure

eliminates toxic reagents and column degradation associated with the reduction

method. Thus, this method is simpler from an instrumental standpoint, exhibits a

higher degree of reagent economy, and is more operator and environmentally friendly.


182

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Chapter 5 – Conclusions and further research


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5.1 Introduction

Eutrophication, which is the enrichment of a body of water with nutrients, may lead to

algal blooms. These events are well recognised causes of aquatic system degradation.

The monitoring of nutrient concentrations in natural waters is an important component

towards improving scientific understanding of eutrophication, as well as providing

information valuable for producing more effective waterway management strategies.

This thesis describes the development and application of flow analysis systems for the

determination of two particular nutrient pools in natural waters, namely total

phosphorus and total nitrogen. The development of a multi-reflective flow-cell using

total internal reflection for application in flow analysis systems is also discussed.

5.2 Total phosphorus

The portable flow analysis system has been successfully applied to the determination

of total phosphorus in situ during shipboard analysis. The analytical characteristics of

the system are; a throughput of 115 measurements per hour, a detection limit of 1.3

µgPL-1 at the 99 % confidence limit, highly linear over the calibration range of 0 –

200 µgPL-1 (r2 = 0.9998), and a precision of 4.6 %RSD at 100 µgPL-1 (n = 10). The

system was deployed aboard the SV Pelican 1 where the acquired data of intensive

temporal and spatial resolution was used for phosphorus mapping. A good degree of

agreement was observed between manual samples and in situ measurements.

However, several problems were identified with the total phosphorus system as

described in Chapter 2:


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1. The reagent storage chamber volume (6 mL) prohibits the instrument from

deployment in a stand-alone fashion for an extended period of time.

Additionally, the single stage gas pressure regulator was found to inadequately

manage the gas pressure inside the reagent chamber, and thus was a cause of

instrumental drift and possible inaccuracy

2. The acidic peroxodisulfate reagent undergoes decomposition in the presence

of heat and acid. After approximately three days, the conversion efficiency of

organic phosphorus compounds began to decrease (Figure 2.14). Additionally,

as this reagent decomposes, acid is generated, which also reduces the

sensitivity of the molybdenum blue photometric method

3. The evolution of bubbles from the acidic digestion environment are a source

of spectrophotometric interference and cause a significant loss of data (29 %

during the shipboard deployment described in Chapter 2.3.7)

4. The hydrolysis of condensed phosphate species is necessarily limited because

the acid concentration required to achieve full mineralisation of these species

on a rapid time-scale (in the order of seconds) also significantly reduces the

formation of phosphomolybdenum blue (Figure 2.7).

The issues described in point 1 can be addressed by increasing the volume of the

reagent storage chamber, as well as installation of a three-stage gas pressure regulator.

Superior regulation of the gas pressure inside the reagent storage chamber will lead to

more reliable measurements.


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The decomposition of peroxodisulfate under acidic conditions (2) can be reduced by

storing the acid and peroxodisulfate separately and merging them in-line prior to

introduction to the sample. Storage under neutral pH conditions should increase the

effective lifetime of the peroxodisulfate reagent.

The debubbler employed in the digestion module (Figure 2.1-3) is effective at

removing bubbles generated during digestion, i.e. prior to collection of the digestate

by the flow analyser. It is likely that bubbles generated, or evolved, during the

analysis are causing the problems described (point 3); in which case, additional

debubbling applied immediately before the detector should decrease the amount of

data loss.

In order to increase the hydrolysis of condensed phosphate species, additional acid is

required. However, this also has the unwanted effect of suppression the formation of

phosphomolybdenum blue (4). This problem can be overcome by merging the

digestate with an alkaline buffer, and thus at least partially neutralising the effect of

any additional acid. However, particular attention to the final reaction pH must be

maintained in order to fully suppress the formation of silicomolybdenum blue.

5.3 The total internal reflective flow-cell

A total internal reflective for cell, consisting of a length of fused silica quartz tubular

capillary, for application in flow analysis system has been successfully constructed

and characterized, with comparison to a conventional z-configuration cell and a

coated capillary multi-reflective cell. This cell was found to have several desirable


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features in common with liquid core waveguides (efficient light throughput that leads

to high signal to noise ratio, versatile choice of irradiant light wavelength) and coated

capillary multi-reflective cells (low hydrodynamic dispersion, no entrapment of

bubbles, high tolerance to refractive index effects).

However, problems encountered during research into the total internal reflective cell

discussed in Chapter 3 include:

1. The total internal reflective cell did not offer any substantial sensitivity

benefits over the comparative z-configuration cell (Table 3.3)

2. The refractive index effect response of the total internal reflective cell was

significantly less than the comparative z-configuration cell, but still more than

that experienced by the coated multi-reflective cell (Figure 3.11)

3. It was not possible to utilise ultra-violet spectroscopy with the total internal

reflective cell, due partly to poor light transmission in that wavelength region

and weak source emission in comparison to that offered by a LED in the

visible region.

The sensitivity of the total internal reflective cell (1) could be improved in two ways.

Firstly, only the light-path that traverses the liquid contributes to the effective

pathlength, and accordingly a capillary with a higher internal diameter to outer

diameter ratio will offer a longer optical pathlength per reflection. This ratio is much

higher for the coated capillary multi-reflective cell (Table 3.2) and is one of the

reasons for its higher sensitivity in comparison to the total internal reflective cell

(Table 3.3.). The second factor is that the angle of light introduction for a total


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internal reflective cell is necessarily limited due to the restrictions imposed by the

critical angle (Equation 3.3). Thus, a capillary material with a higher refractive index

will offer an angle of introduction closer to the normal of the axis of flow, which will

increase the sensitivity offered per unit length of capillary used.

The refractive index effect increases in a multi-reflective cell as the light beam

propagation becomes more longitudinal to the axis of flow. The issues in point 2

could be addressed by choosing a capillary material of higher refractive index and

therefore enabling an angle of introduction more transverse to the axis of flow.

In order to utilise a total internal reflective cell for ultra-violet spectrophotometry (3),

there are three critical requirements: a strong source of ultra-violet light must be used,

a stable and highly ultra-violet transparent coupling material must be available, and a

highly ultra-violet transparent material (such as fused silica quartz) should be used for

the capillary. For the cell described in Chapter 3, both a strong ultra-violet source (30

W deuterium lamp) and a fused silica quartz capillary were readily obtainable;

however, all commercially available coupling materials were either mechanically

stable and not highly ultra-violet transparent (optical glue), or highly transparent but

physically unstable (e.g. water droplet). If a total internal reflective cell is to be

designed to facilitate ultra-violet spectroscopy, then the development of highly

transparent optical glue is essential.

The development of bright far ultra-violet LEDs and materials of increased ultra-

violet transparency in the future will significantly improve and simplify the

development and application of total internal reflective cells in the ultra-violet region.


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5.4 Total nitrogen

A flow analysis system developed for total nitrogen determination has been

successfully applied for measurements in fresh waters. The system has a throughput

of 5 measurements per hour taken in triplicate, a detection limit of 0.05 mgNL-1, high

linearity over the calibration range of 0 - 2 mgNL-1 (r2 = 0.9989), and features a

precision of 1.2 %RSD for 1 mgNL-1 as ammonia (n = 10). Excellent agreement was

found between storm water samples measured using the flow analysis system in

comparison to those obtained using a comparative method.

However, some issues were encountered during the design of the flow analysis

instrument for total nitrogen determination:

1. Operation in estuarine and marine waters was unable to be achieved due to

spectral interference caused by chloride and bromide in the ultra-violet region

(200 - 230 nm) used to determine nitrate (Figure 4.3)

2. A long digestion time (10 minutes) is required to decompose residual

peroxodisulfate in order to limit the suppression of sensitivity caused by its

spectral overlap with nitrate, and thus limits the number of samples that can be

processed to 5 per hour.

In order for the developed total nitrogen method to find application in waters with

high concentrations of chloride (> 100 mgClL-1), such as estuarine and marine waters,

the spectral interference needs to be either eliminated or corrected (point 1). Due to

the large signal caused by chloride in comparison to nitrate, correcting for the spectral


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interference is likely to yield nitrate measurements with a high degree of error.

Chloride (and other interfering ionic species) may be removed by ion exchange;

however, ion exchange columns may have difficulty coping with high ionic strengths

of marine and estuarine waters.

While decreasing the digestion time will be difficult because of the requirement of

achieving 100 % mineralisation as well as decomposition of the residual

peroxodisulfate (2), it may be possible to construct a photo-reactor system with

several reaction chambers that can be used in parallel to perform multiple digestions

simultaneously.

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Publications arising from the research in this

thesis

Publications arising from the research in this thesis

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errata - au-east.erc.monash.edu.au 1.1 a summary of the main components of the nitrogen cycle ......

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