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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
Figure 2.18 Total phosphorus concentration as determined in situ and plotted against
time for 23-Jan-2010................................................................................................84
Figure 2.19 Total phosphorus concentration as determined in situ and plotted against
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
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,
Chapter 1 - Introduction
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].
Chapter 1 - Introduction
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.
Chapter 1 - Introduction
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
Chapter 1 - Introduction
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)
Chapter 1 - Introduction
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
Chapter 1 - Introduction
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].
Chapter 1 - Introduction
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.
Chapter 1 - Introduction
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].
Chapter 1 - Introduction
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
Chapter 1 - Introduction
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
Chapter 1 - Introduction
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
Chapter 1 - Introduction
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]
Chapter 1 - Introduction
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
Chapter 1 - Introduction
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
Chapter 1 - Introduction
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
Chapter 1 - Introduction
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.
Chapter 1 - Introduction
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.
Chapter 1 - Introduction
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.
Chapter 1 - Introduction
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]:
Chapter 1 - Introduction
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
Chapter 1 - Introduction
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
Chapter 1 - Introduction
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,
Chapter 1 - Introduction
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
Chapter 1 - Introduction
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.
Chapter 1 - Introduction
27
1.7 References
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21. Stumm, W., and Morgan, J.J. (1996). Pollution by nitrogen compounds. In
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35. Smith, S.V., Kimmerer, W.J., and Walsh, T.W. (1986). Vertical flux and
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fertilization. Sewage and Industrial Wastes. 24, 925-928.
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45. Estela, J.M., and Cerda, V. (2005). Flow analysis techniques for phosphorus:
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orthophosphate in water. Hydrobiologia 170, 45-49.
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48. Zhang, A., and Oldham, C. (2001). The use of an ultrafiltration technique for
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49. Monbet, P., McKelvie, I.D., and Worsfold, P.J. (2009). Dissolved organic
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injection analyser for laboratory, shipboard and in situ monitoring of nitrate in
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53. Andrew, K.N., Blundell, N.J., Price, D., and Worsfold, P.J. (1994). Flow
injection techniques for water monitoring. Analytical Chemistry 66, 916A-
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instrumentation for monitoring water quality: An Australian perspective.
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57. Jannasch, H.W., Johnson, K.S., and Sakamoto, C.M. (1994). Submersible,
osmotically pumped analyzers for continuous determination of nitrate in situ.
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58. Coetzee, J.F., and Gardner, C.W. (1986). Determination of sulfate,
orthophosphate, and triphosphate ions by flow injection analysis with the lead
ion selective electrode as detector. Analytical Chemistry 58, 608-611.
59. Schubert, F., Renneberg, R., Scheller, F.W., and Kirstein, L. (1984). Plant
tissue hybrid electrode for determination of phosphate and fluoride. Analytical
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61. McKelvie, I.D. (1999). Flow injection analysis. Analytical Testing
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62. Ruzicka, J., and Hansen, E. (1975). Part 1. A new concept of fast continuous
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detector systems. Analytica Chimica Acta 114, 59-70.
72. Tijssen, R. (1980). Axial dispersion and flow phenomena in helically coiled
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techniques and trends. Analytica Chimica Acta 99, 37-76.
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77. Bezerra dos Santos, S.R., Ugulino de Araujo, M.C., and Barbosa, R.A. (2002).
An automated FIA system to determine alcoholic grade in beverages based on
Chapter 1 - Introduction
35
schlieren effect measurements using an LED-photocolorimeter. Analyst 127,
324-327.
78. 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.
79. 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
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265-271.
80. Ellis, P.E., Lyddy-Meaney, A.J., Worsfold, P.J., and McKelvie, I.D. (2003).
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Chapter 1 - Introduction
36
85. Hanrahan, G., Gledhill, M., Fletcher, P.J., and Worsfold, P.J. (2001). High
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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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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)-
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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].
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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):
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
44
S2O82-
(aq) + OH•(aq) � HSO4
-(aq) + SO4
-•(aq) + 1/2O2(g) (2.5)
SO4-•
(aq) + OH•(aq) � HSO4
-(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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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].
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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].
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in 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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
100 !gP/L as phytic acid -
1 min stop time
100 !gP/L as phytic acid -
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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. .
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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)
Flow analysis method Comparative method
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
)
Flow analysis method Comparative method
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
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
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
89
due to bubble evolution remains a confounding factor when measuring natural
samples in the field.
Chapter 2 – A compact portable flow analysis system for the rapid determination of total phosphorus in natural waters
90
2.5 References
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environment. American Scientist 46, 205-221.
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3. Schindler, D.W. (1977). Evolution of phosphorus limitation in lakes. Science
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4. Hodgkin, E.P., and Hamilton, B.H. (1993). Fertilisers and eutrophication in
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quantification and speciation of phosphorus in natural waters. Analytical
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measurement of orthophosphate in shallow wetlands. The Science of the Total
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10. Pérez-Ruiz, T., Martınez-Lozano, C., Tomás, V., and Martın, J. (2001). Flow-
injection spectrofluorimetric determination of dissolved inorganic and organic
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442, 147–153.
11. ANZECC (1992). Australian water quality guidelines for fresh and marine
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and New Zealand Environment and Conservation Council: Canberra.
12. Hara, H., and Kusu, S. (1992). Continuous-flow determination of phosphate
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phosphate after collection on a membrane filter as ion pair of
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L.V. (2000). An atomic absorption spectrometric method for the determination
of phosphorus in foodstuffs using the bismuth phosphomolybdate complex.
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15. Peranemi, S., Vepsalainen, J., Mustalahti, H., and Ahlgren, M. (1992).
Determination of phosphorus in waste water by EDXRF. Fresenius Journal of
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16. Arunachalam, J., John, A., and Gangadharan, S. (1991). Derivative activation
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17. Twinch, A.J., and Breen, C.M. (1982). A comparison of nutrient availability
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18. Lyddy-Meaney, A.J., Ellis, P.S., Worsfold, P.J., Butler, E.C.V., and McKelvie,
I.D. (2002). A compact flow injection analysis system for surface mapping of
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19. Linge, K.L., and Oldham, C.E. (2001). Interference from arsenate when
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27. Lambert, D., and Maher, W. (1995). An evaluation of the efficiency of the
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phosphorus in turbid waters. Water Research 29, 7-9.
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29. Solorzano, L., and Sharp, J.H. (1980). Determination of total dissolved
phosphorus and particulate phosphorus in natural waters. Limnology and
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31. Woo, L., and Maher, W. (1995). Determination of phosphorus in turbid waters
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32. Hinkamp, S., and Schwedt, G. (1990). Determination of total phosphorus in
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33. Benson, R.L., McKelvie, I.D., Hart, B.T., and Hamilton, I.C. (1994).
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34. Golimowski, J., and Golimowska, K. (1996). UV-photooxidation as
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35. McKelvie, I.D., Hart, B.T., Cardwell, T.J., and Cattrall, R.W. (1989).
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36. Solorzano, L., and Strickland, J. (1968). Polyphosphate in seawater. American
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38. Vieira, E.C., and Nogueira, A.R.A. (2004). Orthophosphate, phytate, and total
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Agricultural and Food Chemistry 52, 1800-1803.
39. Thurairatnam, P. (1994). Enzymatic determination of phosphorus in natural
and waste waters. Honours thesis, Monash University, Clayton.
40. Aoyahi, M., Yasumasa, Y., and Nishida, A. (1988). Rapid spectrophotometric
detection of total phosphorus in industrial wastewaters by flow injection
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41. Fernandes, S., Lima, J., and Rangel, A. (2000). Spectrophotometric flow
injection determination of total phosphorus in beer using on-line UV/thermal
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42. Higuchi, K., Tamanouchi, H., and Motomizu, S. (1998). On-line photo-
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43. Sun, F., and Korenaga, T. (1996). Highly sensitive detection system composed
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44. Vlessidis, A.G., Kotti, M.E., and Evmiridis, N.P. (2004). A study for the
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45. Liang, C., Wang, Z.-S., and Bruell, C.J. (2007). Influence of pH on persulfate
oxidation of TCE at ambient temperatures. Chemosphere 66, 106–113.
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48. Peat, D.M.W., McKelvie, I.D., Matthews, G.P., Haygarth, P.M., and
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49. Rumhayati, B. (2007). In situ measurement of phosphorus species in overlying
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57. Motomizu, S., Oshima, M., and Ma, L. (1997). On-site analysis for
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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)
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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.
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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].
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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.
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
112
Research objectives:
• 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.
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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.
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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.
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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.
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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:
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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.
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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.
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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.
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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-
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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.
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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.
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
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.
Chapter 3 – Design and construction of a total internal reflective flow cell for use in flow analysis
133
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pathlength liquid-core waveguide sensor for real-time pCO2 measurements at
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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-
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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,
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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].
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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.
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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.
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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].
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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,
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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)
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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) + OH•(aq) � HSO4
-(aq) + SO4
-•(aq) + 1/2O2(g) (4.10)
SO4-•
(aq) + OH•(aq) � HSO4
-(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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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.
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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.
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
153
Research objectives:
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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.
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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.
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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.
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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).
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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).
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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).
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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.
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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.
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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.
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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.
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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,
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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.
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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.
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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).
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
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.
Chapter 4 – Ultra-violet spectrophotometric flow analysis methods for the determination of nitrate and total nitrogen in freshwaters
182
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Chapter 5 – Conclusions and further research
Chapter 5 – Conclusions and further research
194
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:
Chapter 5 – Conclusions and further research
195
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.
Chapter 5 – Conclusions and further research
196
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
Chapter 5 – Conclusions and further research
197
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
Chapter 5 – Conclusions and further research
198
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.
Chapter 5 – Conclusions and further research
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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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