The Impact of Point Source Pollution on an Urban River, the River

Home , River Medlock, Urban stream

The impact of point source pollution on an urban river, the River Medlock, Greater Manchester A thesis submitted to the University of Manchester for the degree of Doctor of Philosophy in the

Faculty of Science and Engineering

2016

Cecilia Medupin

Table of Contents

The impact of point source pollution on an urban river, the River Medlock, Greater Manchester ...... 1

A THESIS SUBMITTED TO THE UNIVERSITY OF MANCHESTER FOR THE DEGREE OF DOCTOR OF PHILOSOPHY IN THE ...... 1 FACULTY OF SCIENCE AND ENGINEERING ...... 1 2016 ...... 1

Cecilia Medupin ...... 1

Abbreviations ...... 5 Words and meanings ...... 5 Declaration ...... 6 Copyright Notice...... 6 The Author ...... 7 Acknowledgements ...... 8 Abstract 9

Chapter 1 GENERAL INTRODUCTION ...... 10

1.1 COMBINED SEWER OVERFLOWS (CSOS) ...... 12 1.1.1 CSOs and Environmental Regulation ...... 14 1.2 PHYSICAL MODIFICATION ...... 15 1.2.1 Channelisation ...... 16 1.2.2 Examples of modification to the River Medlock ...... 17 1.3 RIVER ECOLOGY: IMPACTS OF PRECIPITATION ...... 19 1.4 WATER POLLUTION ...... 21 1.4.1 Physical and chemical characteristics of rivers...... 21 1.4.2 Biological characteristics of rivers: benthic macroinvertebrates ...... 22 1.5 WATER QUALITY IN THE RIVER MEDLOCK ...... 25 1.5.1 Why study the impact of combined sewer overflows in the River Medlock? ...... 27 1.6 AIMS, OBJECTIVES AND HYPOTHESES ...... 28 1.6.1 Aims ...... 28 1.6.2 Objectives ...... 28 1.6.3 Hypotheses ...... 29 1.7 OVERVIEW AND STRUCTURE OF THE EXPERIMENTAL RESEARCH CHAPTERS ...... 29

Chapter 2 GENERAL METHODOLOGY AND APPROACH ...... 33

2.1 SITE DESCRIPTION OF RIVER MEDLOCK ...... 33 2.2 SAMPLING REGIME ...... 34 2.3 FIELD AND LABORATORY ANALYSIS ...... 38 2.3.1 Benthic invertebrate sampling and analysis...... 43 2.3.1 Benthic invertebrate sampling ...... 43 2.3.2 Spatial and Statistical analysis ...... 44 2.4 WATER QUALITY STANDARDS ...... 46 2.4.1 WFD Standards ...... 46 2.4.2 Classification of Invertebrates...... 47

Chapter 3 LONG-TERM WATER QUALITY OF A HEAVILY URBANISED RIVER: A CASE STUDY OF RIVER MEDLOCK, GREATER MANCHESTER, UK ...... 51

ABSTRACT ...... 51 3.1 INTRODUCTION ...... 52 3.2 METHODOLOGY AND APPROACH ...... 53 3.2.1 Study area ...... 53 3.2.2 Study sites and data collection...... 55 3.2.3 Water quality and ecological parameters ...... 56 3.3 RESULTS ...... 58 3.3.1 Physical and chemical variables ...... 58 3.3.2 Benthic macroinvertebrates ...... 67 3.4 SUMMARY ...... 69 3.5 DISCUSSION ...... 70 3.6. CONCLUSION ...... 72 ACKNOWLEDGEMENTS ...... 73 3.7 REFERENCES ...... 73

Chapter 4 SOURCES OF PO4-P IN AN URBAN RIVER: COMBINED SEWER OVERFLOWS VS WASTEWATER TREATMENT WORKS ...... 78

ABSTRACT ...... 78 4.1. INTRODUCTION ...... 78 Study area: The River Medlock ...... 80 4.2. METHODS ...... 81 4.2.1 Low resolution long term EA data ...... 81 4.2.2 Fortnightly spatial data ...... 84 4.2.3 Data collection ...... 85 4.2.4 High resolution temporal dynamics ...... 86 4.2.5 Data analysis...... 86 4.3 RESULTS ...... 86 4.3.1 Low resolution long term data ...... 87 4.3.2 Fortnightly spatial data ...... 92 4.3.3 Temporal dynamics ...... 96

4.3.4 Comparing PO4-P load and concentration ...... 98 4.4 DISCUSSION ...... 101 CSOs vs WwTW ...... 103 Comparison of phosphorus load in Medlock with other rivers ...... 104 4.5 CONCLUSION ...... 107 ACKNOWLEDGEMENTS ...... 107 4.6 REFERENCES ...... 107

Chapter 5 CATEGORISING THE BENTHIC MACROINVERTEBRATE ASSEMBLAGES AND WATER QUALITY IN A HIGHLY URBANISED RIVER ...... 112

ABSTRACT ...... 112 5.1 INTRODUCTION ...... 113 5.2 METHODOLOGY AND APPROACH ...... 115

5.2.1 Study area ...... 115 5.2.2 Sampling and data collection ...... 116 5.3 RESULTS ...... 127 5.3.1 Physical and chemical variables ...... 127 5.3.2 Benthic macroinvertebrates ...... 135 5.3.3 Relationship between physico-chemical, hydrogeomorphological variables and benthic macroinvertebrate assemblages ...... 143 5.4 SUMMARY OF RESULTS ...... 144 5.5. DISCUSSION ...... 145 5.6 CONCLUSION ...... 149 ACKNOWLEDGEMENTS ...... 150 5.7 REFERENCES ...... 150

Chapter 6 SHORT TERM WATER QUALITY VARIABILITY IN AN URBAN RIVER SUBJECT TO POINT AND DIFFUSE SOURCE POLLUTION ...... 155

ABSTRACT...... 155 6.1 INTRODUCTION ...... 155 6.1.1 Aims and objectives ...... 157 6.1.2 Site Description ...... 157 6.2 METHODOLOGY AND APPROACH ...... 159 6.2.1 Continuous sampling programme ...... 159 6.2.2 Spot sampling ...... 160 6.3 RESULTS ...... 161 6.3.1 Discharge and Precipitation ...... 163 6.3.2 Correlation of physico-chemical variables with discharge and intercorrelation between variables164 6.3.3 Temporal variability ...... 166 6.3.4 Chemical concentration vs discharge ...... 168

6.3.5 PO4-P Load ...... 177 6.4 DISCUSSION ...... 178 6.5 CONCLUSION ...... 180 ACKNOWLEDGEMENT: ...... 181 6.6 REFERENCES ...... 181

Chapter 7 SUMMARY AND CONCLUSIONS ...... 183

7.1 CONCLUSION ...... 187 7.2 GENERAL REFERENCES ...... 187 Appendix ...... 198

Total word count: 47572

Abbreviations 1. ASPT –Average Score Per Taxa 2. ANOVA- Analysis of Variance 3. BIOENV- Biota and Environmental 4. BMWP—Biological Monitoring Working Party 5. BOD—Biochemical Oxygen Demand 6. CSO—Combined Sewer Overflows 7. DO- Dissolved oxygen 8. EA – Environment Agency 9. EQR –Environmental Quality Ratio 10. FFD—Freshwater Fisheries Directive 11. LIFE – Lotic Invertebrate Index for Flow Evaluation 12. WFD-Water Framework Directive 13. SIMPER –Similarity Percentage 14. nMDS-Non-Metric Multidimensional Scaling 15. WwTW – Waste Water Treatment Works 16. WHPT ----Whalley Hawkes, Paisley and Trigg (WHPT)

Words and meanings 1. A combined sewer overflow (CSO) is a collection system of pipes and tunnels

designed to also collect surface runoff, domestic waste water and other waste water

especially during rainfall. It serves as a storage wastewater tank. It is usually available

in houses built before the mid-1960s

2. Hyetograph: Graphical representation of the distribution of rainfall over time.

3. Hydrograph: Graph showing the rate of flow (discharge) versus time

4. WwTW- is a waste water treatment plant where impurities in waste water are removed with the aid of physical structures before the effluent is released to the nearby water bodies.

5. Nutrient load: The quantity of nutrients entering the river in a given period of time

Declaration No portion of the work referred to in the thesis has been submitted in support of an application for another degree or qualification of this or any other university or other institute of learning.

Copyright Notice i. The author of this thesis (including any appendices and/or schedules to this thesis) owns any copyright or related rights in it (the “Copyright”) and s/he has given The University of Manchester certain rights to use such Copyright, including for administrative purposes.

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iii. The ownership of any copyright, patents, designs, trademarks and other intellectual property (the “Intellectual Property”) and any reproductions of copyright works in the thesis, for example graphs and tables (“Reproductions”), which may be described in this thesis, may not be owned by the author and may be owned by third parties. Such Intellectual Property and Reproductions cannot and must not be made available for use without the prior written permission of the owner(s) of the relevant Intellectual Property and/or Reproductions.

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The Author The author has a degree in Biochemistry and a previous experience in industry which led to enrolment for the MSc. Degree in Pollution and Environmental Control at Manchester University. Further experience in industry, environmental regulation and academia, led to an increased interest in integrated environmental science/management and resulted in the author registering for a PhD, the result of which is the current thesis.

Acknowledgements I would like to thank my sponsors, The National Open University of Nigeria for providing the funding for this study.

I would also like to thank my supervisors, Dr. Keith White and Dr. James Rothwell for their continuous help during this research. Thanks to the Environment Agency and to Andy Goodwin, Matthew Harris, Tracey Smith, and Gordon Hardman who provided me with some information pertaining to this study.

I must also thank the people that have assisted me including Marc Attallah who assisted me in the first field survey of high risk sample locations in order to identify combined sewer overflows, walking miles through the Medlock and for all the hard work post field sampling and laboratory analysis. To Dr. Rob Mansfield who assisted me during the project planning, for my colleague Ismael Alkhamaisie, other colleagues who at one time or the other assisted with field sampling- Daryl Teoh, Emma Randle, Irene Okhade, Dr. Amit Bajhaiya, Dr. Merve Engin and a host of others who came out with me on field work. Thanks to Deborah Ashworth, Jonathan Yarwood, Mr. Karl Hennerman, Dr. Gail Challabi and Mr. Graham Bowden of the Geography Department, University of Manchester and Mr. Gary Porteous who analysed my nutrient and metal samples. All the people I have been with at Michael Smith building whose presence provided some fun during my study and to Dr. Andrew Dean, Professor Amanda Bamford for their support.

I acknowledge all the people who have supported me throughout my studies in no small measures: Special gratitude to Dr. Eric Northey and Mrs. Julie Northey and the entire Northey family, Dr. Thomas Keller, Mr. Mark Sullivan, The Manchester Universities’ Catholic Chaplaincy and the Chaplains, Dr. Keith White who provided me the free platform to engage with other aspects of the University in addition to my studies, The Stopford and Michael Smith Building receptionists and security staff. The Methodist International House, the wardens including Mr. Dimitri Brady and all the students and researchers from all over the world who provided me a fun-filled and comfortable house during my studies and writing up of this thesis. Many thanks to Professor Olugbemiro Jegede, Professor Monioluwa Olaniyi, Dr. Felix Olakulehin, and Dr. Seray Ozden. To my brothers and sisters, nieces and nephews and for my parents who have been there for me and to God Almighty to whom I dedicate this thesis.

Abstract The River Medlock is a small (22km) urbanised river, and is one of the five main tributaries which forms part of the River Irwell Catchment in Greater Manchester, UK. The river has a legacy of pollution from the 18th century and continues to be affected by anthropogenic factors including point source pollution from waste water treatment works (WwTWs) and combined sewer overflows (CSOs). In order to investigate the impact of CSOs and the WwTWs on the river hydrology, water quality and ecology of the lower largely urbanised reach, data sets were obtained from the Environment Agency and from direct sampling of the river. Load estimations from continuous discharge records from the river’s gauging station plus estimates of sub-catchment area indicate the lower sites, classified as a “highly modified water body” and downstream of treatment works had had a

higher load of discharge and PO4-P linked to point sources and episodic discharges. Short term, continuous monitoring revealed that CSOs were active during high velocity, but increased concentrations of nutrients post high velocity indicate WwTW effects and possibly diffuse sources. This project reveals that the WwTW

are a major source of PO4-P and that the impact of CSOs on the river quality is short-lived and depends on the degree of precipitation. Other parameters indicate good water quality although the benthic macroinvertebrate community is degraded as a result of episodic increases in the quantity of water destabilising the river bed. Therefore pollution from the CSOs, the WwTW and rapid changes in discharge are the reasons for the river’s failure to conform to WFD requirements.

University of Manchester Cecilia Medupin PhD Environmental Biology Thesis title: The Impact of point source pollution on an urban river, the River Medlock, Greater Manchester

Date: 15th December 2016

Chapter 1 GENERAL INTRODUCTION

Urbanisation is one of the most significant human impacts on the biosphere and exerts a major effect on water quality (Paul & Meyer, 2001) and water resources

(Semadeni-Davies et al. 2008). The impact of urbanisation on fluvial systems results from changes in catchment characteristics due to increased land use for housing, transport, industrial and commercial use (Gregory, 1976). Kulcsaar̂ & White (2012) suggest that urbanisation can proceed in two ways - either in the multiplication of the concentration points (i.e. new conurbations) or in an increase in the size of individual concentrations (i.e. expansion of existing urban areas). The increase in impervious surface cover resulting from urbanisation alters the hydrology and geomorphology of drainage systems and hence the amount of runoff into the receiving water. Rose &

Peters (2001) showed that peak flows were greatest in urban catchment areas, increasing from 30% to more than 100%, compared to non-urbanised catchments.

When this happens, large numbers of properties are at risk of surface water flooding

(Houston et al. 2011). The high concentration of industrial and residential discharges increase nutrient loads plus the amount of heavy metals and other contaminants entering the receiving water course (Clark et al. 2007; Paul & Meyer 2001). Through urbanisation, impervious surfaces such as roofs, roads and car parks replace the natural ecosystems that adsorb pollutants and also degrade organic contaminants.

Consequently, urbanisation promotes the transport of more particulate material and dissolved pollutants from runoff into river systems. Cost projections for supplying water to urban areas are predicted to increase in coming years due to pollution from such diffuse and point sources as well as from population growth (Serageldin 1995).

Income growth per capita will also spur increased agricultural production and this prediction is projected to increase world-wide water pollution by more than 100%

(Serageldin 1995) resulting in a decline in the quality of water entering urban waterbodies from surrounding agricultural areas.

Mullis et al. (1997) found that precipitation intensity and total precipitation volumes strongly influenced both the volume of storm discharges and the duration of storm flows. Such increases in water quantity entering combined sewers systems

(CSS) result in releases of untreated sewage via combined sewer overflows (CSOs).

As a result CSOs have a significant impact on the quality of receiving waters, resulting in the ecological degradation of the watercourse.

Increased rainfall affects surface urban runoff in three ways;

 it increases runoff volume due to reduced rainwater infiltration and

evapotranspiration;

 it increases in the speed of runoff due to hydraulic modification and;

 it decreases the response time of the catchment area - that is the time

between the start of precipitation and the increase in discharge which is

reduced as a result of impervious urban surfaces

These urban rivers are subject to the ‘urban stream syndrome’ (USS), coined to describe the ecological degradation of urbanised water courses (Walsh et al. 2005).

The symptoms of USS include increased impervious surfaces resulting in higher surface runoff velocities and reduced lag time, increased peak discharges (Leopold,

1968), episodic (“flashy”) stream flow (Booth 2005; Roy et al. 2005), greater erosion from altered channel morphology (Walsh et al. 2005; House et al. 1993) and increased hydraulic conveyance efficiency (Goonetilleke et al. 2005). Effects of USS on river quality include increased load of potentially toxic chemicals and organic matter, sediment re-working (Meade 1982) and displacement (‘drift’) of benthic invertebrates

(Lenat & Crawford 1994). Increased water temperatures owing to loss of riparian vegetation and warming of surface runoff on exposed impermeable surfaces are also characteristic of USS (Sinclair Knight Merz 2013). The key sources/causes of urban river pollution are discussed below.

1.1 Combined sewer overflows (CSOs) A combined sewer system (CSS) is a single system that connects the foul and

surface drains directly to the wastewater treatment plant. Combined sewer systems

can cause serious water pollution resulting in the operation of the CSOs when wet

weather flows exceed the sewage carrying capacity of the Wastewater Treatment

works (WwTW). An assessment of the impact of combined sewer overflows (CSOs;

Figure 1-1) on the River Medlock is a key aim of this study. Separate sewers include

surface water or storm water drains and foul drains respectively; CSOs release

dissolved contaminants, inorganic particulates, significant amounts of organic and

suspended solids (Even et al. 2004). The dissolved and particulate organic matter

has a significant effect on oxygenation due to their contribution to the biochemical

oxygen demand (BOD). Deoxygenation damages the biota and impairs the aesthetic

quality of receiving waters, including through the nutrients facilitating growth of

unsightly algae which upon death contribute to the BOD. Various studies have

related poor river quality to CSOs, including Barco et al. (2008), Welker (2008) and

Passerat et al. (2011). They all linked the ‘first flush’ pulse of effluent during intense

rainfall to pollution of the receiving waters.

Figure 1-1: Combined sewer overflow on the Medlock during dry weather (author’s own image)

Under wet weather conditions, CSSs impact the river quality spatially and temporally (Even et al. 2004). Harwood & Saul (2001) describe three components of flow: inflow, continuation flow, and spill flow in a CSO, as shown in Figure 1-1. The impact of CSOs is greatest during wet weather as the CSSs are unable to transport all the wastewater and urban runoff to the WwTW or the capacity of the works is exceeded. Therefore excess flow is diverted to a watercourse (Harwood & Saul, 2001;

Passerat et al., 2011).

Figure 1-2: Schematic diagram of a combined sewer overflow (after Harwood & Saul, 2001).

In the UK, CSOs have long been recognised as one of the major causes of river pollution (Butler & Davies, 2000; Myerscough & Digman, 2008). In 1970, 37% of

14,440 CSOs in England and Wales were reported as unsatisfactory by the Ministry of

Housing and Local Government Technical Committee on Storm Overflows and the

Disposal of Storm Sewage (1970). Some of the problems reported included aesthetic pollution, leading to public complaints, or a marked deterioration in the chemical and biological quality of the receiving watercourse. In 2011, the policy document of the Marine Conservation Society reported over 30,000 CSOs polluted the water courses in the United Kingdom with an uncertified number of discharges to UK water courses (Marine Conservation Society 2011).

1.1.1 CSOs and Environmental Regulation Early work carried out by the then National Rivers Authority (NRA) and the

Foundation for Water Research to assess CSO performance focussed mainly on visual inspection and archived data, as reviewed by Blanksby (2002) and Lau et al. (2005).

They reported that statutory sampling programmes were carried out randomly during working hours and that river monitoring specifically at these times was inadequate to highlight the effect of wet weather flow on the CSOs. Since 1989, CSOs have been regulated by the NRA’s successor the Environment Agency (EA) and they set standards and discharge consents for the commercial water companies who own and manage CSOs (Ayton 1994) such as United Utilities in Northwest England. CSOs are regulated under the European Union’s (EU) Urban Waste Water treatment

Directive (EUUWWTD) (Council of the European Union 1991) and the EU Water

Framework Directive (WFD). The conditions listed in the UK regulation imply that the type, nature of discharge and location of CSOs are critical to their management.

These conditions can be summarised below (Discharge Licenses, personal communication with the EA, 2013).

a. Discharge from CSOs will “only occur as a result of rainfall or snowmelt”

therefore there shall be no discharge in the absence of rainfall.

b. The size of solids in the CSOs should not be greater than 6mm. This is

achieved in the design of mechanically raked screens to prevent heavy floating

solids from entering the water courses.

c. The condition indicates that flows to the storage should only occur when the

flow being passed forward in the sewer is at least 831 Ls-1 (0.831m³sˉ¹). This

implies that the discharge to river shall only occur when the storage is full and

the flow being passed forward in the sewer continues to be at least 831 Ls-1 .

d. The concentration of the discharge should comply with the standard

requirements for effluent released in the river. The CSO discharge should

therefore not contain more than 15mgLˉ¹ of biochemical oxygen demand

(BOD), 6mgL ˉ¹ ammonia and 35 mgL ˉ¹ of suspended solids.

Various studies have reported on the pollution status of the River Medlock since the industrial revolution (Douglas et al. 2002; Burton 2003; Williams et al. 2010; James et al. 2012) arising from point sources. James et al., (2012) suggest that CSOs may be key reasons for many of the rivers of the Irwell catchment (including the Medlock) not achieving the legally required EU water quality standards.

A previous study of the impact of CSOs on the lower Irwell and the upper

Manchester Ship Canal by Rees & White (1993) showed how River Irwell catchment was influenced by storm water overflows, which discharged directly into the river carrying contaminants with a high BOD plus suspended solids, ammonia and PO4-P.

The River Medlock receives episodic discharges from more than fifty CSOs, including

29 within the area under investigation (United Utilities’ ersonal communication,

2013). On the basis of discharge from CSOs, WwTW and river canalisation, the

Environment Agency’s classification of the river using the General Quality

Assessment had been “poor” and has also not met the European Union (EU) Water

Framework Directive which required “good ecological status”.

This study is the first comprehensive assessment of the impact of CSOs on the physico-chemical parameters and their interaction with the benthic macroinvertebrates in the River Medlock.

1.2 Physical modification A water body is regarded as highly modified when its physical characteristics and hydromorphology have been substantially changed e.g. for flood protection and this will have a significant adverse impact on the water use or on the water environment. Globally, changes in land-use practices affect the integrity and conservation of water resources. Changes to the catchment include vegetation removal and deforestation for urban development including housing and industry.

Modification of rivers include culverting, channelisation (to improve transportation), and the creation of dams for irrigation and potable water. These modifications have been shown to alter flow regimes and are regarded as the most serious threat to the

ecological sustainability of rivers and flood plain wetlands (Bunn & Arthington,

2002).

Degradation of the biological components of urban rivers and streams, including the riparian areas, has been documented in a number of studies including

Tong & Chen (2002) who revealed that there was a significant correlation between land use and in-stream water quality on a regional scale in Ohio, USA. Steiger et al.

(2003) reported that the disturbance of riparian vegetation was a major cause of increased sediment loads in rivers. Miserendino et al. (2011), demonstrated that urban sites showed lower biodiversity resulting from domestic sewage inload to rivers and reported large variations in water quality, especially in terms of conductivity, nutrients and dissolved oxygen. High loadings of fine suspended particulates in rivers accumulate in the sediment and affect fish and invertebrates that depend on well-oxygenated habitats (Mainstone et al., 2008 and Heaney et al.

2001).

1.2.1 Channelisation Rivers undergo channelisation either naturally or by human-induced modification (Gregory 2006; Gregory et al. 1992) and the latter result in increased surface runoff in urban catchments (Grimm et al. 2000). The process of re-routing river channels for navigation, reducing erosion (Marshall et al. 1978; Duan et al. 2014) and for flood protection, all contribute to continuous disturbance of the river.

River channelisation significantly impacts on the environment and is regulated through the EU Environmental Impact Assessment Directive (85/337/EEC).

This recognises the potential destruction of the habitat specifically through the disconnection of the river from the flood plain, loss of wetland habitat, silting up of the river and damage to the aquatic ecology. In the UK, the directive is implemented under the Town and Country Planning (Environmental Impact Assessment)

Regulations 2011 for major development works and their impact on natural conservation.

In the British Isles, channel modification began in the 17th century for the purposes of navigation, dredging of existing watercourses to improve drainage and, in the 19th century, to provide gravel for the construction of railway embankments

(Sear and Archer, 1998 in Acreman, 2000). In England and Wales approximately

8504km of rivers have been channelised in response to urbanisation with a channelised density of 0.06 km/km2 and a further 35,500 km of rivers which are regulated (Brookes et al. 1983). Between 1939 and 1945, war-time demand for increased agricultural output and, in later years, the EU’s Common Agricultural

Policy and funding for land drainage improvements, led to intensive and extensive channel modification to ensure continued high rates of agricultural productivity

(Acreman, 2000).

As these physical modifications do not take place uniformly along the river channel (Arnold et al. 1982) they have uneven and specific effects. For example, they directly cause changes in water velocity (Laws & Roth 2004); reduce hydraulic connectivity between river channel and the riparian zone (Laws & Roth 2004); increase the load of suspended solids (Lane et al. 2007); accumulate sediment especially in low-energy river systems (Wilby et al. 1997); and increase contaminant metal fluxes (Longfield & Macklin 1999) associated with soil erosion and fine sediment transport from land (Leemans & Kleidon, 2002). Such modifications have all had their effects on the River Medlock.

1.2.2 Examples of modification to the River Medlock Studies on urban rivers in the Irwell catchment have been described in James et al. (2012). The EA has classified some parts of the River Medlock as a “highly modified water body” based on the degree of modification, in particular the lower

10km reach immediately upstream of the confluence with the River Irwell. Gill’s

(2006) study showed that stretches of most urban rivers in Manchester, including the

Medlock, were culverted as a result of their historical use as sewers, and that

culverting reduces their ability to cope with the large fluctuations in flow rate from precipitation events.

Figure 1-3 shows the River Medlock before and after restoration work in 2014 at Clayton Vale. This stretch of the river was lined with bricks to enhance the flow after serious flooding in 1872 (National Rivers Authority North West 1994). Recent improvements included the removal of brick lining and weirs (Figure 1-4) (James et al. 2012; Manchester City Council 2014). These modifications were aimed at restoring the river to its near-natural state to enhance the invertebrate and fish populations, while at the same time improving the amenity value in areas of public open space

(National Rivers Authority North West 1994).

Before

After

Figure 1-3: Restoration of the River Medlock. View before and after removal of brick lining at Clayton Vale, Manchester. The wall was also removed to allow growth of marginal plants. (Image by Manchester City Council, 2014) NGR: SJ 87359 99209

Before

After

Figure 1-4: Restoration of the River Medlock. View before and following removal of the weir at Clayton Vale (Image by Manchester City Council, 2014) NGR: SJ 88461 99429

1.3 River ecology: Impacts of precipitation River habitats are modified by precipitation-induced changes in flow magnitude and timing which directly influence benthic invertebrate community structure. In addition, changes to the flow regime rework the substrate that also affects the biota (Konrad & Booth 2005). Although river biota has evolved to cope with variations in water velocity (Nilsson & Renöfält 2008), the increased episodicity of urban rivers amplifies the shift in community structure (Willemsen et al. 1990) to 19

favour species capable of withstanding continuous habitat change (Pedersen &

Perkins 1986). Precipitation in urban catchments has been shown to affect water quality (Nilsson & Renöfält 2008; Poff et al. 1997) through the operation of CSOs at high discharge rates (Mulliss et al. 1996).

River hydrographs show increased flood peaks created during storm periods

(Leopold 1968). Such changes in flow pattern in urban rivers are likely to be accentuated by global warming affecting weather patterns (Hulme et al. 2002). For example, high winter flows or short-term increases in discharge from summer storms lead to soil erosion, scouring out of occupied habitats and increased nutrient and suspended solid load which also impact on the biota (Stanley et al. 1994). Low flows significantly modify in-stream communities in lotic systems (Boulton 2003; Lytle &

Poff 2004) by the deposition of silt substrate which exacerbates the impact of other stressors such as high organic pollution and toxins (Boulton 2003) and impact on the river biota. Frequent changes in the sediments, arising from flow-induced deposition and erosion impact on the biota (Whitehead et al. 2009) because changes in bed sediment favour species adapted to unstable habitats such as Chironomidae and

Oligochaeta (Pedersen & Perkins 1986). Higher invertebrate diversity has been observed in stable and coarser sediments due to the increased number of niches

(Collier 1995).

Urban rivers in Manchester including the Medlock catchment (Figure 1-5) are subject to heavy rain and prolonged periods of wet weather conditions with over

1,000 mm of rain per year and short periods of high precipitation reaching 0.11 mm hr-1 (National River Flow Archive). One of the objectives of this study was to monitor the river during wet and dry weather conditions. The aim was to assess the impact of flow on water quality and the benthic invertebrate biota.

1.4 Water Pollution Point and diffuse source pollution include biodegradable organic material, metals and nutrients. These result in changes in physico-chemical characteristics of rivers, in particular temperature, inorganic and organic suspended solids, nutrients, dissolved oxygen, biochemical oxygen demand (BOD), conductivity and pH. Levels of trace metals can increase as a result of pollution. Changes in discharge and water velocity can also be due to human impact as described above.

1.4.1 Physical and chemical characteristics of rivers River temperature is a key physical parameter which influences river ecology

(Webb & Walsh 2004) and it is in turn influenced by altitude and source. In Britain, lowland river temperature can become as warm as 25°C, while at high altitudes, streams remain cool at round 11oC, which is the national average (Hynes 1960; Orr et al. 2010). Temperature has an indirect influence on the mobilisation as well as the toxicity of pollutants. For example (Li et al. 2013) and Doudoroff & Katz (2014) found that increased temperatures enhanced the release of phosphorus from sediments.

Temperature also influences the distribution and abundance of benthic macroinvertebrates due to interspecific differences in thermal tolerance(Quinn &

Hickey 1990; Leunda et al. 2009). Total Suspended Solids (TSS) represents the actual measure of mineral and organic particles transported in the water column by mass.

TSS can originate from sewage pollution, soil erosion, agricultural activities or industrial runoff and algal blooms. Resuspension of silt, sand, clay or gravel from the interstices of the river bed will also contribute to the suspended solid load depending on water velocity (Everest et al. 1987). Concentration of ammonia-N and 5-day biochemical oxygen demand (BOD5) are used to determine the impact of organic pollution on water quality and the animal community (Donald et al. 2002) Ammonia-

N is released by the largely microbial-mediated breakdown of organic material while

BOD is an indirect measure of the amount of biodegradable organic material (Drury et al. 2013).

Trace metals, including cadmium, chromium, copper, mercury, nickel, lead and zinc are ubiquitous in nature (Gasperi et al. 2012) and hence are found in rivers due to erosion. As a result of the anthropogenic factors described above, mobilisation and transport of trace metals are increased by storm events and flooding (Rose et al., 2004). An example of trace metal pollution of an urban catchment area is the River Irwell. Work by Eyres & Pugh-Thomas (1978), indicated a reduced number of benthic invertebrates due to metal pollution, plus the presence of pollution tolerant taxa such as Chironomidae and Oligochaeta.

1.4.2 Biological characteristics of rivers: benthic macroinvertebrates Elliott et al. (1980) define macroinvertebrates as those organisms that are retained by a net or sieve with an aperture of 0.6mm. However a larger aperture, generally as 0.95mm, is more commonly used by the research community and regulatory bodies such as the UK’s Environment Agency although such a size will not retain the early stages of some aquatic insects.

The dissolved oxygen (DO) content of a river affects the types of invertebrates.

Many invertebrates such as Plecoptera and Ephemeroptera require well aerated water while some such as Oligochaetaes- (commonly of the family Tubificidae) are abundant and the diptera Chironomidae are very abundant in poorly oxygenated water. Both feed on detritus, including anthropogenic sources (Butcher et al. 1927).

Tubificidae feed on detritus in the sediment and many Chironomidae feed on bacteria and detritus on the sediment surface. A rise in temperature affects the amount of oxygen saturation, especially in turbulent streams where there is ready exchange with the atmosphere and hence the water is fully oxygenated.

The nature of a river bed, which can be eroding or depositional, also determines the types of invertebrates to be found. Most invertebrates show structural adaptations which enable them to live either in fast or slow flowing water. Mayfly nymphs, of the family Ecdyoduridae (synonym Heptageniidae), are flattened with clawed appendages and apply themselves close to stones to reduce resistance to flow. 22

Many case-bearing caddisflies make their cases of stones to increase their density.

The caseless caddis larvae use silken threads for anchoring themselves, and the larva of the black fly Simulium, spins a small mat of silken thread and attaches itself to this via a complex circlet of tiny hooks (Hynes 1970).

Feeding habits among benthic macroinvertebrates vary and include carnivores, such as many large stoneflies, some caddisflies (e.g. Rhyacophilidae), beetle larva (Coleoptera) and leeches (Hirudinea). Most of the nymphs of mayflies and many caddisflies, scrape (graze) algae off stones (Moon, 1939). Detritus of terrestrial (allochthonous) vegetable origin is fed on by many organisms of the shredder and collector guilds such as some stonefly nymphs (Hynes 1960) and caddisfly larvae (Percival & Whitehead 1929). Food is also carried away by currents, including small particles of detritus, detached algae and small benthic invertebrates that have lost their attachment. This is then available to animals downstream such as the blackfly Simulium, which filters the water by making continuous grasping movements with a pair of mouth brushes.

The assessment of water quality using benthic macroinvertebrates has several advantages over physical and chemical analyses as it not only gives an indication of the quality of the water for living organisms but is also able to detect and integrate environmental change that could only be directly assessed through continuous monitoring of physico-chemical parameters. Hellawell (1986) argues that benthic macroinvertebrate analysis has the following benefits:

 they are the most useful indicators for monitoring water quality because they

are sensitive to toxic pollutants and the general degradation of the river,

 they are reasonably sedentary and are therefore representative of local

conditions (Cook 1976),

 different taxa display different levels of sensitivity to organic and other

pollutants and their responses are well understood,

 they have lifespans long enough to provide a record of environmental quality

(Pratt & Coler 1976),

 they are relatively large and easily identified,

 they are ubiquitous in freshwaters and their biogeography is similar

throughout the world.

There are various ways to assess the resident macroinvertebrate community in a river. These include passive sampling, including the use of colonisation samplers which are regarded as a passive method by Hellawell (1978); and active sampling, using, for example, hand-held nets. Colonisation samples have the advantage of eliminating differences arising from changes in substrate and hence facilitate inter- site comparisons; however, they may not reflect the indigenous community at a given site. Various samplers that are used for invertebrates include traps (drift and emergence traps); colonisation samplers (multiplate); and immediate samplers (grab, air lift, corer, dredge) (Elliott et al. 1980). The period of exposure of these traps and colonisation samplers in freshwater varies from four to six weeks (Weber 1973). The main advantage of active sampling is that the animals are collected at the same time as the sampling is conducted and that it can be performed on a range of substrates.

Various biological indices have been used to indicate pollution and reflect the differing sensitivity of freshwater organisms to organic pollution, including the

Biological Monitoring Working Party (BMWP) (Hawkes 1997) and its variant the

Whalley, Hawkes, Paisley & Trigg (WHPT) metric, (Paisley et al. 2014; WFD-UK

Technical Advisory Group (UKTAG) 2014). Other workers have assessed indices based on relative abundance, including MacNeil et al., (2002) who proposed the

Gammarus:Asellus ratio. Ephemeroptera, Plecoptera and Trichoptera (EPT) are taxa widely used in combination as bio-indicators as they are relatively intolerant of organic pollution (Hoiland et al. 1994; Malmqvist 2002; Ode et al. 2005).

1.5 Water Quality in the River Medlock Various studies on the Medlock showed it has been polluted since at least the mid-nineteenth century due to its industrial legacy (Burton 2003; MacKillop 2012;

Williams et al. 2010) and urban catchment (Willey 2011) resulting in diffuse sources of pollution, WwTWs and localised storm events which had led to frequent spill events from combined sewer overflows (CSOs) (Rees & White 1993; James et al. 2012;

Douglas et al. 2002; Frost et al. 1976; Tyson & Foster 1996 & Environment Agency

2009c).

Previously, the Medlock in common with other UK rivers was classified based on the General Quality Assessment (GQA) and General Quality Assessment

Headline Indicators (GQAHI). Data obtained from 1990 to 2009 from the source to

Lumb Brook 12km downstream showed the river was rated by the EA as “very good”; from Lumb Brook for a further 6.5km to the confluence of the River Irwell the river was rated from “poor” to “fairly good”. This latter section is classified by the EA as a “heavily modified river” based on channel modification. A similar spatial pattern was observed for nutrients between 1990 and 2009. While NO3-N concentrations in the river were generally “very low” from the source to Lumb

Brook, the concentration recorded from Lumb Brook to the confluence with the

River Irwell 6.5km downstream were higher and therefore were classed

“moderate”. For the same period, PO4-P concentration was “high” at the upper sites

(with a range of 0.2mgP Lˉ¹) and the sites below the WwTW, 0.5km below Lumb

Brook had “very high” concentrations (1mgPLˉ¹). The EA therefore attributed the greatest challenge to good water quality on the Medlock to the “very high” concentration of PO4-P from the WwTWs plus CSOs (EA, personal communication,

2013).

The River Medlock is typical of many rivers in the UK in that it has a mixed use catchment; in this case 40% urbanised in the lower reaches with agricultural areas (26%) and woodland (18.8%) above. In common with many northern UK catchments, the agriculture in the upper Medlock catchment (Figure 1-5) is largely

sheep and cattle pasture with little arable farming (Farming and Countryside

Education (FACE) 2007). The River Medlock, like other urban rivers, is subject to multiple stressors (Heathwaite 2010) including poor water quality arising from point sources, in particular WwTWs and CSOs. Diffuse pollution within the urbanised reach includes road runoff, badly connected sewers, old landfills, and agriculture (James et al. 2012). Re-engineering of the Medlock, although not as extensive as in some other urban rivers such as the Irwell (Williams et al., 2010) will also influence flow and hence pollutant behaviour and ecology, including the benthic invertebrate community.

Figure 1-5: River Medlock catchment boundary with urban settlements. The sampling sites (S1 to S6). (Source: ArcGIS)

Several studies have looked at the impact of point source pollution on urban rivers from various single angles. But this study provides a holistic understanding on the basis of water quality (physico-chemical parameters); biology (benthic macroinvertebrates), nutrient dynamics and the impact of precipitation on water quality. The catchment of the Medlock contributes to local understanding of the River

Catchment Development Project which feeds into the national, EU Water Framework

Directive and global water management plans (Sustainable Development Goals).

Thus, the solution following this study will be sensitive to local needs as it provides information necessary to control pollution and its challenges at a local scale. Also, it will inform effective decision-making processes and practical solutions, in this case, the control of discharge entering the river.

1.5.1 Why study the impact of combined sewer overflows in the River

Medlock?

CSOs have been seen as major point sources of pollution of urban rivers after the WwTWs (James et al. 2012). Although both sources are controlled through legislative requirements, discharges from CSOs are less regulated and also storm dependent. Thus, while the WFD requires good ecological status, many rivers have yet to comply with these standards. Therefore, in establishing the magnitude of CSO impact, river managers will be able to focus efforts on controlling high discharging

CSOs. In addition, the assessment of CSO impact on the river relative to WwTWs would assist water managers and environmental regulators to implement cost- effective and focussed mitigation strategies to facilitate compliance with the WFD.

One of the aims of this study is therefore to identify and quantify the main source(s) of PO4-P in the Medlock by determining its concentration at different seasons and to assess the points on the river with the highest concentration. In combination with knowledge of discharge it will then be possible to quantify amounts of PO4-P both from WwTWs and from CSOs.

1.6 Aims, Objectives and Hypotheses

1.6.1 Aims The overall aim is to examine the impact of point-sources, specifically WwTWs and

CSOs, on the water quality and benthic invertebrate community of the River

Medlock. More specifically, this study aims to:

1. Characterise the chemical and biological status of the River Medlock based on

the requirements of the WFD. This includes long term (10 years) changes in

the physicochemical condition and the benthic macroinvertebrate community.

2. Assess the relative contribution of CSOs and WwTWs on the water quality of

River Medlock. This will include the relative importance of these point sources

on the load of selected pollutants, including PO4-P, and during episodic

rainfall events.

3. To characterise the benthic macroinvertebrate community in the river and to

assess the relative importance of the physical, chemical and

hydromorphological impacts on the community.

1.6.2 Objectives The above aims give rise to the following objectives.

1. To assess past and current (2000-2013 and 2013-2014) water quality status by

measurements of key physico-chemical and hydromorphological parameters

indicative of anthropogenic-induced change, including re-engineering of the

river and the degree of urbanisation.

2. Estimate fluvial WwTW NO3-N, and PO4-P load and quantify the relative

importance of other sources, in particular CSOs.

3. To determine the role of short-term (15min- over 92 days plus spot sampling at

intervals corresponding to rainfall events) discharge dynamics on water quality.

4. To determine the diversity and abundance of the benthic macroinvertebrates

community and the relationship with the above environmental variables.

5. To use the data obtained over the sampling period to assess the extent of

compliance of the River Medlock with the EU WFD standards. This will include

an examination of the relative importance of chemical quality and

hydromorphological parameters in defining the status of the benthic

macroinvertebrate community. Such information will aid in the formulation of a

strategy for the future management of the Medlock and other urban rivers.

Long-term changes where examined from data provided by the EA and the

current (2013-2014) water quality and ecological status was assessed from

sampling and analysis at a number of points along the river.

1.6.3 Hypotheses 1. CSOs contribute to the poor water quality and reduction in the benthic

macroinvertebrate community in the River Medlock.

2. Control of CSOs will reduce the pollution of the River Medlock rather than

further improvements in the WwTW effluent.

3. “Good” water quality as defined by the WFD does not result in “Good

Ecological Status”.

Although it is appreciated that the recent decision to withdraw from the EU will have implications for water management, including compliance with the WFD, EU standards will still apply until withdrawal, and possibly thereafter (Miller 2016).

1.7 Overview and Structure of the Experimental Research Chapters The experimental chapters 3 to 5 are written in the form of papers. Each chapter therefore has an abstract, introduction, aims and objectives, methods, results, discussion and conclusion.

Prior to the experimental chapters, chapter 2 provides a more detailed account of the methodology and approach used by the EA, and in this study, to assess water quality and benthic invertebrate biota. It includes a description of the study area, the rationale for site selection and the parameters analysed during the period 2013-2014 seasonal survey and the short term examination of discharge and water quality

dynamics. Analytical methods and statistical analytical tools are described. The EA’s

General Quality Assessment and the WFD requirements for rivers are outlined.

Chapter 3 is an overview of the studied reach of the river based on the datasets between 2000 and 2013 obtained from the EA prior to my study between 2013 and

2014. One study site was located upstream and two downstream of the main

Wastewater Treatment Works (WwTW). The major tributary to the study area, Lord’s

Brook was also analysed to assess its pollution impact to the river. The study indicated that high PO4-P concentration was a major barrier to the river’s compliance with the WFD. The benthic invertebrate community failed to achieve the “good ecological status” required by the WFD as it was dominated by pollution-tolerant taxa even though water chemistry, other than PO4-P, indicated good water quality.

High suspended solid concentrations greater than the EU Freshwater Fisheries

Directive requirement of 25mgL¯1 were recorded at certain periods.

Subsequent chapters 4 to 6 determined through further interrogation of the EA datasets plus sampling and analysis by myself at six sites, between 2013 and 2014. In chapter 4, the source, dynamics, load and relative contribution of PO4-P in relation to river episodic conditions was determined: the long term EA data, annual PO4-P concentrations from the fortnightly data at the six sites from 2013 to 2014 and the high resolution dataset were obtained from August 2014 to 31st October. These dataset provided a detailed account in time and space of the PO4-P dynamics in the river. Although the PO4-P load estimated (Webb et al. 1997) was less than 3.5 kgha⁻¹yr⁻¹ during any of the study periods, it was within the range found in other urban areas and agricultural sites. The treatment works contributed an average of

92% of PO4-P load which made it a major contributor, rather than the CSOs or other diffuse sources.

In Chapter 5, the environmental variables which created the greatest impact on the benthic macroinvertebrate community in the Medlock were identified using the annual fortnightly datasets and a combination of biotic indices and multivariate tools. The results indicate the river invertebrates were influenced by a number of

hydrogeomorphological and chemical factors, mainly discharge, altitude, slope and,

PO4-P. Altitude, slope, were linked to site location and hence reflects discharge and flow. The apparent relationship between the invertebrate community and PO4-P concentration was attributed to the relationship with discharge. It is also suggested that the community was impacted by the river’s sandy substrate (<2mm) which are not suitable for pollution-sensitive taxa. The catchment of the uppermost site was only 33% urbanised compared to the other sites, which had >40% urban catchment.

Given that streamflow patterns are modified by urban development it is suggested that hydrological changes impact on the biota. The new biotic indices introduced by the WFD in 2015 i.e. Whalley Hawkes, Paisley Trigg (WHPT) scores and WHPT

Average score per taxon (WHPT ASPT) were used in the assessment alongside the old BMWP score. Both indices gave a very similar indication of the degree of environmental degradation at all sites. It is concluded that the key stressor that degrades the invertebrate community of the Medlock is urban stormwater runoff released to the river by the hydraulically efficient drainage system that responds rapidly to changes in precipitation. Other stressors, such as pollution (rather than discharge) from CSOs plus WwTWs are considered of less importance. Therefore the

Medlock can be considered to be suffering from the “urban stream syndrome”

(Walsh et al., 2005) which adversely affects the distribution and abundance of benthic macroinvertebrates in urbanised rivers. Until the symptoms of the urban stream syndrome are addressed the Medlock is unlikely to comply with the WFD.

Chapter 6 employed the higher resolution continuous datasets taken at the

EA Gauging station to examine short-term changes in water quantity and quality arising from the hydraulically efficient and hence ‘flashy’ nature of the urbanised lower reaches of the Medlock. This part of the study showed that the CSOs were sometimes active during short-term high rainfall events although other sources, probably agricultural and road runoff contributed PO4-P and suspended solids to the river during such events. The study therefore confirmed that CSOs were not the sole pollution sources as revealed by the EA for some highlighted peak 31

discharge periods (EA personal communication, 2015). The results of this chapter supported the earlier studies as to the importance of discharge on river sediment destabilisation, and hence on the benthic macroinvertebrate community. This chapter is informative for policy makers and the water companies as they tend to focus on PO4-P reduction, removal and possibly, recovery from the WwTW.

However, dealing with phosphorus and other pollutants will not fully address the reasons for the degraded invertebrate community due to the urban stream syndrome. Therefore suggestions for reduction of discharge into the river are included. Chapter 7 provides a summary of the chapters and a general conclusion from the study.

This comprehensive study provides a clear understanding of the issues within the River Medlock catchment. It also serves as a pilot study that will help to integrate projects which will address catchment restoration plans in a cost- effective way. The study therefore contributes effectively to knowledge of river catchment studies.

Chapter 2 GENERAL METHODOLOGY AND APPROACH

2.1 Site Description of River Medlock The River Medlock (Figure 2-1) is a third-order stream in one of the heavily urban areas of Greater Manchester. The river forms part of the River Irwell catchment which in turn is a component of the Mersey catchment and which is one of the largest in the UK. The Medlock is sourced from the moorland to the north east of Oldham (National Grid Reference: SD 95308 05431) where it flows for 22km through Ashton-under-Lyne and continues in a south easterly direction to discharge into the River Irwell immediately downstream of Manchester city centre

(SJ 85781 97858).

The Medlock has a catchment area of approximately 57km2 and about 40% of the land cover is urbanised (National Rivers Flow Archive, 2016); the remainder is recreational or agricultural land (Tyson & Foster 1996). The main tributaries of the

River Medlock are Thornley, Taunton, Glodwick, Lumb Clough and Lord’s Brooks.

For the last 10 km, the River Medlock flows through the Manchester city centre mainly in underground culverts and artificial channels until it confluences with the

River Irwell.

The river has a continuously operational waste water treatment works (WwTW) at Failsworth (NGR: SJ 89674 99800), fifty combined sewer overflows (CSOs)

(personal communication, United Utilities, 2013) and an unknown number of surface water drains within the study area (

Figure 2-2). Failsworth WwTWs is situated 12.6km south of the river’s source.

The daily discharge from this WwTW was obtained from the water utility company, United Utilities. Table 2-1 provides estimated values of the frequency/volume of discharge from the CSOs in 2013 (Personal Communication,

United Utilities, 2013). Population served by WwTW is 21,624 (United Utilities’ personal communication, 2016).

Figure 2-1: The River Medlock showing the EA sample sites, EA gauging station, CSOs (graduated values),the WwTWs and the urbanised areas of the catchment. Dashed line shows the source to Lumb Brook & from Lumb Brook to the confluence with the River Irwell; see introduction (Section 1.5). CSO Data from United Utilities (Source: ArcGIS.)

Table 2-1: Combined sewer overflows (CSOs) on the River Medlock classified on the basis of number of spill events, duration and volume of discharge (United Utilities, 2013)

S1&S2 S3 S4 S5 S6

No of CSO events 219.1 112.4 420.1 72.1 1 (spills/year) Duration (hours) 603.3 171.5 3547.9 79.9 0.6 Volume (m3) 1,029,584 52,945 585,608 29,902 1,005

2.2 Sampling regime Long term datasets from 2000 to 2013 were obtained from the EA. Physico- chemical parameters were available for three sites on the river: here designated S0

(Medlock Vale), S4 (Millstream Lane) and S6 (Pin Mill Brow) and, for the tributary

Lord’s Brook from 2000 to 2006. Complete physico-chemical datasets were available for S6, while S0 and S4 were monitored by the EA between 2000-2004 and 2010-2012;

hence no data is available between 2005 and 2009. The variables analysed were pH, dissolved oxygen, temperature, conductivity, suspended solids, biochemical oxygen demand (BOD), ammonia-N, NO3-N and PO4-P. Benthic macroinvertebrates were assessed by the EA for S6 in autumn and spring, and at the Lord’s Brook between the period of 2000 and 2008.

Annual sampling was carried out by the author on the river from March 2013 to

April 2014 at six locations (S1to S6). These sites were selected upstream and downstream of twenty-nine CSOs and the main WwTW. Two of the sites, S4 and S6 corresponded to the EA sites (long- term sampling) while the EA site S0 was not sampled in this study due to poor access. Sampling frequency was fortnightly for physico-chemical variables and monthly for benthic invertebrates. On each sampling date, samples were taken from three sections of each site to obtain an average for the measured parameters. S3 was not sampled for benthic macroinvertebrates due to poor access and S6 was sampled less frequently due to safety considerations at certain periods, in particular during high flows.

Table 2-2 describes the sample location, sub-catchment area, and distance from the source, velocity, depth and width of the river.

Figure 2-2 shows the position of the sampling sites.

Table 2-2: Sample locations on the River Medlock, sub-catchment area, distance from source, and geographical variables. Site 0 was only sampled by the EA

Catchment Distance Catchme area as % of (km) Altitude Slope Site name Site No. nt area Latitude Longitude the total from (m) (%) (km²) catchment source Mill Brow S1 15 26 6.60 53.5173 -2.0892 138.51 2.69 Park Bridge S2 23.5 41 8.50 53.51282 -2.0997 117.87 2.34 Road Daisy Nook S3 29.7 52 10.30 53.50107 -2.12398 88.81 2.21 Garden Millstream S4 43.9 76 13.00 53.49258 -2.16317 66.81 1.92 Lane Purslow Close S5 53.7 93 16.10 53.48197 -2.21164 47.35 1.67 Pin Mill Brow S6 54.4 95 17.40 53.47726 -2.21571 42.39 1.57 WwTW n/a n/a n/a 12.60 n/a n/a n/a n/a Lord's Brook 4.5 2.58 12.07 53.49422 -2.15519 68.16 -

Medlock vale S0 65 37.2 12.23 53.49311 -2.15135 74.82 -

Figure 2-2: Study sites (S1-S6), Lord’s Brook, Environment Agency’s Gauging station and the WwTw. Dashed line shows the source to Lumb Brook & from Lumb Brook to the confluence with the River Irwell; see introduction (Section 1.5). Blue circles represent estimates of volumes of discharge from CSOs entering into the River Medlock. (Source: ArcGIS).

S1 S2

NGR: SD 94183 NGR: SD 93489

02262 01798

S3 S4

NGR: SD 91874 NGR: SJ 89272

00493 99554

S5 S6

NGR: SJ 86052 NGR: SJ 85781 98382 97858 Figure 2-3: Photographs of sample sites S1 to S6 (shown on the map of

Figure 2-2). S6 shows the debris screen which is aimed to retain large objects and prevent flood damage.

The study sites had riparian vegetation, largely erosional at S1-S3 and depositional downstream of the river, hence the large silty substrate observed downstream at S6.

Invertebrate colonisation samplers were installed in the river at S2 and S6 for four-months from September to December 2014.

Higher resolution sampling was carried out at the Environment Agency’s gauging station site from 1st August 2014 to 31st October 2014. Monitoring of pH, dissolved oxygen, temperature, conductivity and turbidity at 15-minute intervals was carried out. Fifteen minute continuous discharge records from the gauging station was obtained from the EA for the duration of sampling.

The summary of the sampling regime carried out on the river for the study is provided on Table 2-3.

Table 2-3: Summary of the measurements carried out on the River Medlock. * Spot samples collected during this period.

Data from the This study: This study: high Environment seasonal resolution analysis Agency analysis Duration 2000-2012 March 2013 - 1st August 2014 - April 2014 31st October 2014 Sampling Regime Monthly Fortnightly 15-minute continuous - high resolution Temperature √ √ √ Dissolved Oxygen √ √ √ pH √ √ √ Conductivity √ √ √ Nutrients NO3-N, PO4-P, NO3-N, PO4-P, *NO3-N, *PO4-P, * ammonia-N TP, ammonia-N ammonia-N BOD √ √ - Suspended solids √ √ √ Turbidity - - √ Benthic S6 Spring/Autumn Monthly at S1- 4 months invertebrates between 2000 & S2, S4 -S6 2008

2.3 Field and laboratory analysis Sampling of physico-chemical parameters and benthic macroinvertebrates were obtained at the river. Photographs of sites S1 to S6 are shown in Figure 2-3.

The historical data obtained directly from the EA used standard analytical methods as described in the Standard Committee of Analysts Publications, (2011).

For the seasonal analysis between March 2013 and April 2014, measurements of pH, dissolved oxygen (percentage saturation and mgL-1), temperature and conductivity were obtained using a pre-calibrated hand-held multiparameter water quality meter (YSi 556 Multi probe system YSI, Yellow Springs, Ohio, USA).

Fortnightly discharge was calculated for each sub-catchment areas on the basis of their relationship to the total catchment area. Discharge records obtained at the continuously monitored EA gauging station was considered a preferable option to in- situ measurements because discharge based on estimates of velocity and cross- sections area was not always possible or accurate due to limited safe access, particularly at periods of high velocity and discharge. Therefore, the data presented in Table 2-4 are approximations. Continuous discharge data was available only at the

EA’s Gauging station 0.5km below S6. The river discharge measured in cubic metre per second (m³s-¹) for each study location was obtained by estimation using a simple linear regression equation which correlated the catchment area (km2) with mean discharge (Q) for twenty-eight rivers within Greater Manchester including the River

Medlock (National Rivers Flow Archive, 2014). The model y = 0.0218x - 0.0422, which indicated a strong correlation between the catchment areas and river discharge (R² =

0.873) was used to estimate the discharge at the sites in relation to the sub-catchment areas.

Water velocity was measured using the float method i.e. by recording the time taken for the float to travel over a given distance (10m) along the river. The results obtained from this exercise was compared with the EA velocity records for the same period. River substrate class was obtained by estimating the percentage substrate observed at each site and is shown on Table 2-5. Substrate composition of benthic habitats was determined at each sample site by visual examination of the percentage coverage of each particle class: silt, sand, gravel, pebble, stones, boulder, and 39

bedrock. Substrate composition was divided into the above classes based on the modified Wentworth scale (Cummins 1962).

Table 2-4: Values for average (no. = 23) of Width (W) (m), minimum and maximum depth (D) (m), Velocity (V) (msˉ¹) and discharge (Q) (msˉ¹) at sample sites.

W D V Q

Sites Ave Ave Min Max Ave V Min V Max V Ave Q Min Q Max Q W D D D (msˉ¹) (msˉ¹) (msˉ¹) (m³sˉ¹) (m³sˉ¹) (m³sˉ¹) (m) (m) (m) (m)

S1 5.7 0.22 0.13 0.3 0.27 0.16 0.53 0.15 0.05 0.46

S2 8.2 0.23 0.12 0.59 0.26 0.11 0.59 0.23 0.08 0.72

S3 8.6 0.44 0.28 0.55 0.27 0.13 0.53 0.29 0.1 0.9

S4 8.8 0.27 0.14 0.64 0.6 0.14 1.25 0.43 0.15 1.34

S5 8.8 0.29 0.11 0.69 0.57 0.13 1 0.53 0.18 1.63

S6 8.5 0.29 0.15 0.58 0.57 0.13 1 0.53 0.18 1.66

One litre water sample was collected in acid-washed (10% hydrochloric acid) polypropylene bottles. A 300ml aliquot was filtered through a dried (heated to 500°C for three hours) pre-weighed 0.45 µm glass microfiber filter (Whatman GF/C filter,

VWR International, Leicestershire UK) to remove any organic particulates. The filter paper was then oven dried at 105°C for 24 hours to remove moisture, weighed andthe difference in weight taken to determine the total suspended solids (SS) in the filtered sample.

Part of the aliquot containing inorganic constituents was filtered through a

0.45µm Millipore (Millipore Limited, UK) for measurement of ammonia (mgL-1 as ammonia-N), NO3 (mgL-1 as nitrate-N), PO4 (mgL-1 as phosphate-P) and trace metals

(µgL-1) while 40ml unfiltered sample was preserved for the analysis of total phosphorus by acid digestion. The trace metal samples were acidified with two drops of ultra-pure reagent nitric acid to pH~2 to retain the metals in solution.

The nutrients NO3-N and PO4-P were analysed using a SEAL Auto Analyzer 3

High Resolution instrument (SEAL Analytical Ltd, Southampton, UK). This equipment has a high level of precision and ultra-low detection limits (SEAL

Analytical, 2013). Detection limit for phosphate measured as P was 0.004 mgLˉ¹ and nitrate measured as N was 0.05 mgLˉ¹. For further information on the methods employed by the auto analyser see SEAL Analytical (2013). Phosphorus is measured by the Environment Agency as soluble reactive phosphate, again as the element P in mgL-1. Total phosphorus (a measure of the total inorganic and organic phosphorus) was processed through a pressure digestion technique with sulphuric acid and potassium persulphate (Mackereth et al. 1978) and analysed using ion chromatography.

Ammonia was analysed using the Hanna low range reagents kit (HI-93700-01;

Hanna Instruments Ltd, Leighton Buzzard, UK) and the colour change quantified by spectrophotometry at an absorbance of 500nm. The limit of detection for ammonia measured as N was 0.01 mgLˉ¹.

For measurement of Biochemical oxygen demand (BOD), a brown glass bottle was used for the collection of samples from each site to avoid autotrophic metabolism and incubated at 20°C for five days. The five-day BOD was calculated as the difference between dissolved oxygen at day zero and day five measured using a pre- calibrated Hanna dissolved oxygen meter (Hanna Instruments Ltd, Leighton

Buzzard, UK).

The trace metals chromium, cadmium, copper, nickel, lead and zinc were analysed by Inductively Coupled Plasma-Mass Spectrometry (ICP-MS) using an

Agilent 7500cx (Agilent Technologies, Santa Clara, USA) spectrometer. Calibration was by matrix-matched standards.

slopes were obtained by dividing each site’s elevation from the river’s source by the distance of the sample site from the source (See Table 2-2).

To obtain the high resolution temporal dynamics discharge, a 15-minute duration discharge record was obtained from stage records supplied by the EA’s gauging station SJ849975 at London Road, Manchester. Precipitation data of 15- minute duration was obtained from the Whitworth Meteorological Observatory,

Manchester.

Spot samples in parallel with the high resolution discharge data were collected for measurement of PO4-P, NO3-N, ammonia-N, and suspended solids, from the

London Road gauging station. The samples were obtained over a range of low and high flow conditions between August and October 2014 in order to establish concentration-discharge relationships. Samples were obtained by lowering a bucket into the river from the bridge located immediately upstream of the gauging station.

These samples were then decanted into a one-litre sample container for subsequent analysis.

Table 2-5: Mean values for types of substrate found at sample sites examined at different times. The text in bold shows the dominant substrates to be stones and sand, which constitutes 60% of the total substrate composition recorded in the river.

Boulders Stones Pebbles Gravel Site Sand (%) Silt (%) Mud (%) (%) (%) (%) (%)

S1 6 37 7 5 30 15 3.5 S2 10 39 7 7 27 12 3 S3

S4 4 27 1 23 32 12 3 S5 14 43 3 6 28 8 0 S6 6 18 1.4 3.8 42 16 2 % total 8 33 4 9 32 13 2 composition

2.3.1 Benthic invertebrate sampling and analysis All macro invertebrate samples collected through kick sampling and from artificial colonisation samplers were preserved in 70% ethanol and counted in the laboratory (Pawley, Dobson, & Fletcher, 2014). All macroinvertebrate specimens apart from Oligochaeta were identified to family level using the taxonomic groups used in the biological monitoring working party (BMWP) score.

2.3.1 Benthic invertebrate sampling

2.3.1.1 Kick-net sampling Samples were collected using a 1mm mesh hand net by the three-minute kick net sampling technique outlined in the Water Framework Directive, UK policy report

(UK Technical Advisory Group 2008). A one-minute manual search was also carried out to collect benthic invertebrates that could have been missed through kick sampling.

2.3.1.2 Inter-site comparisons using colonisation samplers Colonisation samplers allow a comparison between sites as colonisation is independent of the natural substrate (Czerniawska-Kusza 2004). Colonisation samplers were therefore applied in this study to facilitate examination of the impact of stream hydrology and substrate on the invertebrate community.

Sampler Description The polypropylene pall ring colonisation samplers were based on the S.Auf.U colonisation samplers described by Watton & Hawkes (1984). The pall rings were strapped together in one position using plastic ties to form the colonisation sampler.

The samplers were attached at the base to a one-mm mesh size white polyester netting to trap the benthic macro-invertebrates which are defined as those that are retained on a 1mm mesh (Tagliapietra & Sigovini 2010) (Figure 2-4). The samplers were anchored to two house bricks joined together (21 cm x 20 cm). They served to

weigh down the sampler in the river which was attached to a rope. The surface area available for colonisation was 0.042m².

Figure 2-4: Invertebrate colonisation sampler mounted on a brick

Sampling strategy Two artificial colonisation samplers were positioned on the river bottom at sites S2 and S6 (see Figure 2-4). Site S2 had lower concentration of suspended solids, low discharge and nutrient concentration. Also, S2 was upstream of the main operational WwTWs and the substrate was composed of a mixture of stones, gravel and sand. Site S6 had a higher concentration of suspended solids, higher nutrient concentration, had a higher discharge and is located below the WwTWs. Site S6 had a sandier substrate compared to S2. The colonisation samplers were left in the river for a four month period from September 2014 to December 2014 and removed at 30 day intervals. The 30-day period is suitable for the development of a representative community of organisms (Weber 1973; Meier et al. 1979).

2.3.2 Spatial and Statistical analysis The sub-catchment areas were determined using the package Terrain Analysis

System, GIS (Lindsay 2005) and a 50m (horizontal) and 0.1m (vertical) digital terrain model (Centre for Ecology and Hydrology DTM). Statistical analysis were carried out

using Microsoft Excel 2013, SPSS (IBM, 2013), GraphPad Prism version 6 and

PRIMER 6 (Clarke & Warwick 2001). Differences between the physico-chemical variables and the biotic indices (BMWP, ASPT) at each site were analysed using One- way Analysis of Variance (ANOVA) on the assumption of data homogeneity and normal distribution for each category of independent variable Pearson correlation analysis was used to investigate how the various metrics changed and how these variables were related to pollution impact

Principal component analysis (PCA) based on a correlation matrix between samples was used to analyse physico-chemical variables. The first few PCs allow an accurate representation of the true relationship between the samples in the original high dimensional space as summarised by the percentage variation explained (Eigen values). All the datasets were standardised in order to obtain comparable scales

(Clarke & Warwick 2001).

Non-metric multidimensional scaling (nMDS) were used to test for similarities in benthic invertebrate abundance. The computation of similarity was performed using the triangular matrix generated through Bray-Curtis similarity test (Clarke &

Warwick 2001).

Differences in benthic macroinvertebrate composition among sample sites and between the sample seasons were analysed using the similarity percentages

(SIMPER) routine. The abundance data was square root transformed in order to give more weight to abundance in comparison species. This was followed by the Bray-

Curtis similarity test which was employed among all pairs of samples and seasons to describe assemblage similarity (Clarke & Warwick 2001).

The BIOENV procedure (Clarke & Warwick 2001) was used to select environmental variables (EVs) that best explained benthic invertebrate community patterns by maximising a rank correlation between their respective resemblance matrices. Data were square root transformed and normalised to allow comparison at

the same scale. The weighed Spearman rank correlation coefficient (ρ) between the physico-chemical and benthic invertebrate community similarity matrices was the basis for this procedure. The physico-chemical with the largest ρ was taken to identify the best match with the benthic invertebrates. Multivariate analysis were performed using the PRIMER-6 software package (Clarke & Warwick 2001),

Statistical package for social sciences (SPSS) for the analysis of variance (ANOVA).

Particle size- flow distribution: The particle size distribution and the flow conditions under which sediment was eroded, transported or deposited, was estimated using the Hjulström-Sundborg Diagram (Figure 2-5) described by Earle

(2015).

Figure 2-5: The Hjulström-Sundborg diagram showing the relationship between particle size and flow velocity. It shows the tendency of the sediments to be eroded, transported or deposited. (Source: Earle,2015).

2.4 Water Quality Standards

2.4.1 WFD Standards Under the Water Framework Directive, river chemistry is classified as shown in Table 2-6 and includes altitude of the river as a surrogate variable for river gradient and hence one of the natural characteristics that might influence ecological

communities. In addition the biota of fast flowing and hence high altitude rivers are more vulnerable to organic pollution and this is reflected in the different standards for rivers above and below 80m (UKTAG Water Framework Directive 2013). The

WFD classifies PO4-P from “High” to “Poor” for altitude either less than or greater than 80m (WFD-UK Technical Advisory Group, 2012). River Medlock has an altitude between 31 and 376m AOD according to the National Rivers Flow Archive.

Table 2-6: Classification of river chemistry according to the WFD at 90%ile

Variables High Good Moderate Poor BOD (mgLˉ¹) (Altitude < 80m) 4.0 5.0 6.5 9.0 BOD (mgLˉ¹) (Altitude > 80m) 3.0 4.0 6.0 7.5 Ammonia-N (mgLˉ¹) 0.6 0.6 1.1 2.5 (Altitude < 80m) Ammonia-N (mgLˉ¹) 0.2 0.3 0.75 1.1 (Altitude > 80m) DO (% saturation) >80 79 64 50

PO4-P (mgLˉ¹) 0.05 0.12 0.25 1.0 (Altitude 80m)

2.4.2 Classification of Invertebrates The Biological Monitoring Working Party Score (BMWP) is a method used in the assessment of rivers with reference to the freshwater aquatic families (except oligochaetes which are identified to class) based on the taxonomy of Maitland (1977) and BMWP (1978). Families that are very sensitive to sewage pollution receive scores of 10 and the most tolerant families receive a score of 1. Thus, the sum of the total scores from the samples collected will determine the category to which the river is classified as shown in Table 2-7. The average score per taxon (ASPT) is the ratio of the

BMWP Score to the number of scoring taxa (N): BMWP Score/N and is less dependent on sampling effort. In this study, the BMWP score and Average score per taxa (ASPT) were calculated in each instance to allow comparison with earlier work.

Table 2-7: BMWP Score, ASPT and interpretation (Hawkes 1997)

BMWP Score ASPT Category Interpretation 0-10 ≤3.9 Very poor Heavily polluted 11-40 4.0 - 4.9 Poor Polluted or impacted 41-70 5.0 -5.9 Moderate Moderately impacted 71-100 6.0 - 6.9 Good Clean but slightly impacted >100 > 9 Very Good Unpolluted/unimpacted

However, the BMWP which was formerly used for water quality status classification is being replaced with a new measure called the Whalley, Hawkes,

Paisley & Trigg (WHPT) metric which aligns with the requirements (Article 8; Section

1.3 of Annex II and Annex V) of the WFD (2000/60/EC). The River Invertebrate

Classification Tool (RICT) (Paisley et al. 2014) is a model used to contextualize WHPT scores by using a model to predict site-specific reference values and provide a WFD compliant probabilistic classification. WHPT was designed to detect organic enrichment as well as other stressors to the invertebrates Therefore WHPT relates the response of invertebrate taxa to organic enrichment using a ‘pressure sensitivity score’ (PSs). The PSs is the sum of the PSs assigned to each taxon present in a single sample from a single season (WFD-UTAG, 2008).

There are two differences between the original BMWP and the new WHPT.

The first is that WHPT considers numerical abundance (Paisley et al. 2014) as shown on Table 2-8. Therefore increasingly abundant intolerant taxa attract a higher score

(e.g. Perlidae: AB1, 12.6; AB4, 13.0) and the reverse is the case for low scoring tolerant taxa (e.g. Glossiphonidae: AB1, 3.4; AB4, 0.8). Secondly, the BMWP is based on analysis of 82 taxa whereas the WHPT is based on 106 taxa so its sensitivity is slightly greater.

Table 2-8: WHPT (Whalley & Hawkes, Paisley and & Trigg) logarithmic abundance categories

Abundance category Numerical Abundance AB1 1-9 AB2 10 – 99 AB3 100 – 999 AB4 >1000

The new metric comprises of WHPT NTAXA (sum of the number of different taxa contributing to the assessment from the same sampling site) and WHPT ASPT

(average score per taxon) which is applied as an abundance weighted metric AB (AB

= abundance related pressure sensitivity score for each taxon contributing to the assessment). Both metrics are assessed separately and then combined in a “worst of” approach to provide the overall invertebrate classification.

The WHPT ASPT is applied as abundance weighted metric (Table 2-8) calculated as WHPT Classification: Count/Abundance category/Score; WHPT ASPT =

Sum AB/WHPT NTAXA. Therefore, observed value of ASPT = PSs ÷ NTAXA. The observed value is then converted to bias-corrected values. Bias correction is estimated for the value of ASPT for taxa missed because of sample sorting and identification errors by using the equation: Estimated ASPT of missed taxa = 4.29 + 0.077x observed value of NTAXA where the observed value of NTAXA is the value prior to bias correction.

Therefore, in order to determine the biological status of the river based on the

WFD criteria, the ASPT of the samples observed (Obs) is divided by the predicted

(Pred) pristine condition score using the RICT statistical model with the WFD setting.

The RICT model therefore provides a classification Ecological Quality Ratio (EQR) and an estimate of the probability of the result belonging to any of the WFD classes as shown in Table 2-9 for both metrics based on observed data (UKTAG), 2014). EQR =

Observed/Predicted; EQR values close to one therefore indicate invertebrate communities close to the natural state, those near to zero indicate a high level of pollution or disturbance. 49

Table 2-9: EQR for WHPT-ASPT and WHPT-NTAXA

WHPT ASPT- EQR NTAXA EQR High/Good 0.97 0.80 Good/Moderate 0.87 0.68 Moderate/Poor 0.72 0.56 Poor/Bad 0.59 0.47

Chapter 3 LONG-TERM WATER QUALITY OF A HEAVILY URBANISED RIVER: A CASE STUDY OF RIVER MEDLOCK, GREATER MANCHESTER, UK

Abstract This paper examines the water quality and ecology of an urbanised 5km reach of the River Medlock catchment for over a decade between 2000 and 2013. The aim was to identify the main challenges to achieving good water quality, including compliance with European Union Directives. Dataset were obtained from the

Environment Agency for physico-chemical parameters and benthic macroinvertebrates. Three locations were examined: one upstream and one immediately downstream of the single operational wastewater treatment works

(WwTW), plus a third 5km further downstream and 6.5km above the confluence with the River Irwell. The tributary Lord’s Brook which brackets the first two sites was assessed to determine its status and impact on the Medlock. The WwTW was the major source of PO4-P and although concentrations reduced with time remains much higher than the 0.1mgPLˉ¹ standard indicative of good water quality. Other variables including ammonia and BOD were generally within standard requirements for EU Rivers. Despite the generally good water quality, biotic indices

(BMWP, ASPT and EQR) indicate the river to be moderately polluted which suggests the impact of non-sewage related pollution or some other stressors. On the basis of the WFD standard for PO4-P and ecological quality, the river remains polluted and has not markedly improved over the period 2000-2013.

Key words: River Medlock, urban water quality, waste water treatment works, benthic macroinvertebrates

3.1 Introduction The River Medlock in Greater Manchester, UK is a mixed use catchment with two-thirds of the catchment being classified as “highly modified” by the UK

Environment Agency (EA). The modified section of the river starts from one of the river’s tributaries, Lumb Brook until its confluence 6.5km downstream with the River

Irwell. The lower part of the Medlock drains a highly urbanised catchment and receives effluent from a major waste water treatment work (WwTW).

For 200 years, the rivers of Greater Manchester, including the Medlock, deteriorated in quality due to increases in industrialisation and urbanisation. Serious flooding in the mid-Nineteenth Century led to canalisation, culverting, and installation of weirs. These activities increase the flow of flood water and are likely to damage the ecology of the river (Williams et al. 2010).

De-industrialisation and improvements to wastewater treatment resulted in some recovery and according to a 2007 report by Manchester City Council (2007) river quality had improved considerably, including reduced sewage contamination and nutrient concentrations compared to the 1990s. Water quality improvements resulted from improvements to the WwTWs plus a reduction in NO3-N following improved farming and agricultural practices (Environment Agency, 2007; European

Union 2010). However, the continued high PO4-P concentration in the river is linked to the discharge from the WwTW and CSOs as reported by The Irwell Catchment

Pilot Steering Group (James et al. 2012). High PO4-P and the effect of physical modifications degrading the benthic invertebrate community are the major reasons why the river had not met the requirements of the Water Framework Directive

(WFD)(Council of the European Union 2000; European Environment Agency 2015).

To manage a river with a history of pollution such as the Medlock and hence to achieve compliance with the WFD requires knowledge of long term water quality plus qualitative and quantitative information on point and diffuse sources of

pollution. Until now, no published study has been carried out to establish the long- term changes in river quality of the River Medlock.

The WFD requires all waters to reach ‘good ecological status’ by 2027 so the results from this study can contribute to identifying reasons for non-compliance by the Medlock. The overall aim of this study was to assess the efficacy of the water management measures on the river’s quality over time from EA’s long term water quality dataset and suggest potential improvements.

Objectives

1. Using EA datasets, assess the long-term water quality dynamics and

determine the abundance and diversity of the benthic macroinvertebrates

community.

2. To determine the major source of pollution to the river.

3. To determine the status of the river with respect to the Water Framework

Directive (WFD).

3.2 Methodology and Approach

3.2.1 Study area The River Medlock (Figure 3-1) rises in the Pennines that surround Strinesdale to the north east of Oldham in Greater Manchester (National Grid Reference: SD

95308 05431). The river has a catchment area of 57.5km2 and the major tributary within the study area is Lord’s Brook. The Medlock catchment is heavily urbanised

(40%), including light industry that extends from the south Pennine to Manchester

(CEH, 2014). The Medlock flows for 22km through steep sided woodlands and continues in a south westerly direction to discharge into the River Irwell immediately downstream of Manchester city centre (SJ 85781 97858). The river receives episodic discharges from more than fifty CSOs (United Utilities, personal communication,

2014).

The river has two WwTWs: a major treatment works at Failsworth (SJ8982

9979) and a smaller treatment works at Park Bridge (SD 9394 0253). The Failsworth

WwTW operates primary, secondary (biological filters) and tertiary (Nitrifying trickling filters) treatments while the treatment works at Park Bridge (SD 9394 0253), operates solely by the secondary rotating biological contactors. While, the

Environment Agency permits the Failsworth WwTW to discharge effluent volume limit of 16,000 cubic metres per day, with a maximum dry weather flow of 6180 m³d¯¹, the Park Bridge WwTW has a markedly lower discharge permit volume limit of 20 cubic metres per day at a rate of 0.0007m³s¯¹ (0.7 Ld¯¹) (Environment Agency, personal communication, 2014).

Figure 3-1: River Medlock, Greater Manchester showing catchment and urban settlements. Insert: Map of the UK with location of Greater Manchester. (Source: ESRI, GIS)

3.2.2 Study sites and data collection The EA river sites were selected based on their relative proximity to the main

WwTW, the location of CSOs on the river and the extent of datasets collected from such locations. The study sites (as shown on Figure 2-1) are S0 - Medlock Vale

(SJ9032899673); S4-Millstream Lane (SJ8909799436) and S6- Pin Mill Brow

(SJ8567297756). The major tributary along these three study sites -- Lord’s Brook

(SJ8987999773) is 0.8km upstream of the main WwTW was also included in the assessment. The EA collected samples from the river monthly, with varied sampling frequencies at each site between the period 2000 and 2013.

Site 0 (S0) is 3km above the main WwTW, 8km downstream of a high discharging CSO plus others discharging less frequently and the small WwTW at

Park Bridge (SD 9394 0253). It is located within the Medlock Vale Park.

Site 4 (S4) is located 0.6km downstream of the Failsworth WwTW and hence receives effluent from the WwTW, and also from Lord’s Brook. The river also runs through an urbanised area for 0.1km and then Clayton Vale local nature reserve. In

2014, the brick lining at Clayton vale, created after the flooding in 1872, was removed together with weirs and other barriers as part of the EA project of restoring the river back to its natural state (Manchester City Council 2014).

Site 6 (S6) is located 5km downstream of the Failsworth WwTW within a highly urbanised area and, because of its shallow slope collects silt and debris. The

EA therefore installed a debris screen to collect debris and hence reduce impediments to flow during flood conditions. The river flows underneath the A6010 in Manchester city centre before the confluence with the River Irwell.

In order to identify CSO infrastructure discharging into the River Medlock, effluent discharge licences were obtained from the EA in 2013. This information was complemented by the water companies’ provision of specific points of discharge in the study areas, simulation data on spill analysis, including frequency of spills per

year, duration and volume of CSO discharge per year. The information extracted for the study showed that twenty-nine CSOs were located within the study areas.

The sub-catchment area for each sampled location, their sizes relative to the

EA’s gauging station and their distance from the river’s source at Strinesdale reservoir is summarised on Table 3-1.

Table 3-1: River Medlock catchment information

Station Name Station Catchment Catchment Distance Discharge Altitude No. area (km²) area as % of (km) from (m³s¯¹) (m) total source Lord’s Brook 2.58 4.5 12.07 0.01 68.16 Medlock Vale S0 37.2 65 12.23 0.77 74.82 Millstream Lane S4 43.9 76 13.04 0.91 66.81 Pin Mill Brow S6 54.4 95 17.40 1.14 42.39

3.2.3 Water quality and ecological parameters Monthly physico-chemical datasets of biochemical oxygen demand (BOD), suspended solids, conductivity, temperature, dissolved oxygen, nitrate (as NO3-N) and phosphate (as PO4-P) were obtained from the Environment Agency for the three sites and tributary for the following periods: S0 and S4 from 2000 to 2004 and 2010 to

2013 (no data from 2005 to 2009) while S6 had a complete dataset from 2000 to 2013.

Datasets for Lord’s Brook was obtained from 2000 to 2006 and, had been included in the study in order to assess its contribution to pollution of the Medlock and receives discharge from CSOs. The water quality was analysed in accordance with the

Standard Committee of Analysts Publications (EA Standard Committee of Analysts

Publications, 2011) and APHA (1989). Benthic macroinvertebrates were sampled by the EA twice yearly during spring and autumn. Benthic invertebrate data was only available at S6, and only between 2000 and 2008 and for Lord’s Brook between 2000 and 2006. Benthic macroinvertebrates were identified to family level with the exception of Oligochaeta which was identified to class in accordance with the requirements of the UK’s Biological Monitoring Working Party (BMWP) score

(Wright, Moss, Armitage, & Furse, 1984). A summary of the EA sampling regime is shown on Table 3-2.

The assessment of the macroinvertebrates based on the spring and autumn seasons was carried out to apply the results to the River Invertebrate Classification

Tool (RICT) in order to assess the environmental quality of the river. This model compared observed information to an expected “pristine” condition expected of the river and classified on the basis of environmental quality bands from A = “Very good” to F = ”Very bad”. The river is therefore classified with the Environmental

Quality Index on the basis of the BMWP score and the ASPT of the sample obtained for the two seasons.

Table 3-2: Datasets obtained from the EA

Date/site S0 S4 S6 Lord's Brook 2000-2004 √ √ √ √ 2005-2009 n/a n/a √ 2005-2006 2010-2013 √ √ √ n/a Dissolved oxygen (% saturation) √ √ √ √ Suspended solids (mgLˉ¹) √ √ √ √ Conductivity (µScmˉ¹) √ √ √ √ NO3-N (mgLˉ¹) √ √ √ √ PO4-P (mgLˉ¹) √ √ √ √ √ (between 2000 Benthic macroinvertebrates n/a n/a √ and 2008)

Continuous discharge data was available only at the EA’s Gauging station

0.5km below S6. Therefore river discharge was measured in cubic metre per second

(m³s-¹) for each study location and this was obtained by estimation using a simple linear regression equation which correlated the catchment area (km2) with mean discharge (Q) for twenty-eight rivers within Greater Manchester including the River

Medlock (National Rivers Flow Archive, 2014). The linear regression equation y =

0.0218x - 0.0422 which indicated a strong correlation between the catchment areas and river discharge (R² = 0.873) was used to then estimate the discharge at the sites in relation to the sub-catchment areas of S0, S4, S6 and Lord’s Brook. Discharge was determined in order to estimate the nutrient load entering the river at each sub- catchment area.

The ratio of mean NO3-N to mean PO4-P concentration was determined for all data obtained from the EA in order to determine the impact of the Waste water

Treatment Works (WwTW).

3.3 Results

3.3.1 Physical and chemical variables

3.3.1.1 River Medlock Figure 3-2A-H show the pattern of dissolved oxygen, conductivity, suspended solids, BOD, ammonia-N, NO3-N and PO4-P at sites S0, S4 and S6. The mean and standard error of the mean were calculated for the three sites on the basis of available data i.e. from 2000 to 2004 and from 2010 to 2013 for S0 and S4 and from 2000 to 2013 for S6. pH had an overall mean of 7.9 + 0.3 (SD) and was within the normal range for rivers (WFD-UK Technical Advisory Group (UKTAG) 2012) and hence not suggestive of pollution. Temperature had an overall mean of 10.37+ 0.64. pH and temperature were not analysed further since there was no difference between sites or over time.

Over the 2000-2013 period of study, the average discharge of the Medlock was

0.73m³sˉ¹ (Figure 3-2 A). A one-way analysis of variance indicated there was no significant (p >0.05) difference with year, but larger variations were found in 2000 and

2008 due to the higher than average rainfall in February, from September to

November in 2000 and, in January, and from July to November in 2008.

Between 2000 and 2004, & 2010 and 2013, higher dissolved oxygen (>90%) was recorded in the river (Figure 3-2B), indicating very good level of oxygenation. While there was no difference between the sites for the average conductivity, suspended solids, BOD and ammonia-N (Figure 3-2C-F) in the river, a significant (p<0.05) difference between the three sites was found for NO3-N and PO4-P concentration

(Figure 3-2G-H, Table 3-3). The peak concentration (mgLˉ¹) observed for BOD, ammonia-N and suspended solids in 2001 (25/9/2001) at S6, suggested a pollution

incident. Communication with the EA revealed that a significant pollution incident had been reported to the EA two days earlier on the 23/09/2001. The pollution was caused by discharge of organic chemicals/products into the river about three miles upstream of S4 (NGR: SD 90699 03183). Following this date, readings were found to be within the normal range of concentration by the EA samplers (EA, personal communication 2016). Elevated concentration of suspended solids recorded at the three study sites in 2001 (18/10/2001) and 2012, indicates the influence of high precipitation and therefore elevated discharge levels (EA, personal communication

2016). The highest total precipitation recorded for these elevated concentrations occurred in October, 2001 and during January and April, 2012.

NO3-N and PO4-P were highest at S4 and S6. One-way analysis of variance

(ANOVA. post Hoc, LSD) showed a significantly (p<0.05) lower concentration at S0 for the periods assessed (Table 3-3). Between 2000 and 2013, S6 showed a decline in the concentration of PO4-P from 0.77 mgLˉ¹ to 0.60mgLˉ¹, which indicated “poor” quality (Table 3-4). Although NO3-N was lower than the recommended General

Quality Assessment, the values of PO4-P were higher than the recommended level for good ecological status of 0.1PO4-PmgLˉ¹ (European Directives 91/676/EEC;

91/271/EEC; 96/61/EEC; 2000/60/EC). A peak in ammonia-N was observed in 2001 for S6 (Figure 3-2F) with mean concentration recorded at 1.26mgLˉ¹. This was classified as “poor”. Although BOD values reached 5.23mgLˉ¹ (Figure 3-2E) for the same period and location, it was within the WFD requirement.

All other measured variables complied with the WFD standards (UKTAG

Water Framework Directive 2013) during the period of measurement. However, the concentration of suspended solids was higher than the recommended 25 mgLˉ¹ of the EU Freshwater Fisheries Directive (78/659/EEC & 2004/44/EC).

C D

E F Altitude > 80m Altitude > 80m

Altitude > Altitude <80m 80m

G H

Figure 3-2: Mean± SEM annual water quality at the three sites S0, S4 and S6 from monthly samples on the River Medlock, 2000-2004, 2005-2009 and 2010-2013. A. Discharge; B. DO; C. Conductivity; D.

Suspended solids; E. BOD; F. Ammonia-N; G. NO3-N; H. PO4-P. The dotted lines B to H represent WFD standard requirement for “good ecological status” for surface water quality, Freshwater 60

Fisheries Directive for Suspended solids (D). In some cases the WFD standards vary with altitude 80 m AOD.

Table 3-3: One way ANOVA to compare S0, S4 and S6 for variables measured from 2000 to 2004 and from 2010 to 2013. No. of samples at S1= 9, S2 = 9; S3=14

2000 - 2004 2010 - 2013 S/No Variables Comments (S0, S4 & S6) (S0, S4 & S6) Dissolved F2,12 =11.32, F2,9 = 0.77, 1 High DO levels oxygen (%) p < .05 p > .05 Conductivity F2,12 = 0.721, F2,9 = 0.09, No difference 2 (μScmˉ¹) p > .05 p > .05 between sites Suspended F2,12 = 0.50, F2,9 = 0.09, No difference 3 solids p > .05 p > .05 between sites (mgL¯¹) F2,12 = 1.85, F2,9 = 3.71, No difference 4 BOD (mgL¯¹) p > .05 p > .05 between sites No difference between sites Ammonia-N F2,12 = 1.56, F2,9 = 5.93, between 2000 and 5 (mgL¯¹) p > .05 p < .05 2004. Higher at S6 between 2010 and 2013 NO3-N F2,12 = 32.40, F2,9 = 22.69, Highly different at 6 (mgL¯¹) p < .01 p < .01 the sites PO4-P F2,12 = 111.27, p F2,9 = 28.97, Highly different at 7 (mgL¯¹) < .01 p < .01 the sites

Table 3-4: Pearson correlation of variables with time at S6 from 2000 to 2013. Number of samples at S6 = 14

S/No Variables 1 Dissolved oxygen (%) p < .05 r = 0.57 2 Conductivity (μScmˉ¹) p < .05 r = 0.86 3 Suspended solids (mgL¯¹) p > .05 r = - 0.02 4 BOD (mgL¯¹) p > .05 r = - 0.30 5 Ammonia -N (mgL¯¹) p > .05 r = - 0.42

6 NO3-N (mgL¯¹) p > .05 r = 0.84

7 PO4-P (mgL¯¹) p < .05 r = - 0.72

The substrate (Figure 3-4) recorded during the spring and autumn seasons at S6, between 2000 and 2008 showed that the river was largely composed of pebbles

(≤64mm, 33% to 48%), gravel (≥2mm; 30% to 32%) and sand (≤ 1mm, 19% to 35%).

6 0 S p rin g A u tu m n

t a

4 0

t s

b 2 0

S %

P C S u b s tra te T y p e s

Figure 3-3: Types of substrate found at the River Medlock at S6 during spring and autumn seasons for 2000 and 2008.

3.3.1.2 Lord’s Brook Over the period 2000 and 2006, the dissolved oxygen levels recorded at the

Brook were high (>80%), as in the river. Similarly, Brook pH and temperature were similar to the records on the river. Although conductivity declined from a maximum value of 600µScm¯¹ in 2001 to 468µScm¯¹ in 2004, the levels recorded were similar to that in the river (Figure 3-4A).

Average concentration of suspended solids complied with the standards of the

Freshwater Fisheries Directive of 25 mgLˉ¹. However, a peak value of 38 mgLˉ¹ was recorded in year 2000 and linked to above average precipitation as found in the river

(Figure 3-4B). All other parameters analysed at Lord’s Brook were lower than the

WFD requirements, including BOD and ammonia-N (Figure 3-4 C & D); the exception was PO4-P (Figure 3-4F) and which recorded the highest concentration in 62

2005 (0.79mgLˉ¹). There was no reported incident to the EA to account for the peak concentrations of BOD and ammonia recorded in 2002. (EA, personal communication

2016).

Figure 3-4: Mean (± SEM) annual water chemistry parameters at Lord’s Brook between 2000 and

2006. (A) conductivity; (B) suspended solids; (C) BOD; (D) Ammonia-N; (E) NO3-N; (F) PO4-P. The dotted horizontal line represents the WFD standards concentration on the basis of altitude, except suspended solids which is based on the EU Freshwater Fisheries Directive.

3.3.1.3 Relationship between NO3-N and PO4-P concentrations

A correlation between of NO3-N and PO4-P often indicates the WwTWs as a contamination source (Jarvie et al. 1998). A correlation analysis of NO3-N and PO4-P data indicated a significant (p<0.05) positive correlation at S4 with r = 0.791, n =100 and S6 (r = 0.679, n =168) while there was no correlation at S0 (Figure 3-5). This analysis suggests that the WwTW is the likely source of PO4-P discharge to the two downstream sites.

S 4 3 .0 S 0 0 .6

2 .5

) 1

- 2 .0

)

1 g

- 0 .4

( g

1 .5

(

O 4

P 1 .0 O

P 0 .2

0 .5

0 .0 0 .0 0 1 2 3 4 5 0 2 4 6 8 1 0 1 2 1 4

- 1 - 1 N O -N (m g L ) N O -N (m g L ) 3 3

3 .0 S 6

2 .5 )

1 2 .0

m (

1 .5

4 O

P 1 .0

0 .5

0 .0 0 2 4 6 8 1 0 1 2 1 4 1 6 - 1 N O -N (m g L ) 3

Figure 3-5: Relationship between average concentration of PO4-P (mgL¯¹) and average NO3-N (mgL¯¹) at site S1 and S2 (from 2000 to 2004; 2010 to 2013) and from S3 (2000 to 2013).

The average NO3-N / PO4-P ratio at the three sites further indicated the influence of the point sources on the river as shown on Table 3-5 and,

Table 3-6 for Lord’s Brook. The highest N/P ratio was found at S0 (>20) while

S4, S6 and Lord’s Brook had ratios which were less than 11 for the periods. N/P ratios which are typically less than 11 confirm the influence of the WwTW and as found elsewhere by Jarvie et al. (1998). It indicated periods when the nitrogen was limiting resulting from greater availability of the PO4-P from sewage inputs, followed by the influence of urban and industrial activities. The low ratio recorded at Lord’s Brook suggested the influence of episodic pollution from CSOs since the brook was located above the treatment works.

Table 3-5: Comparison of average concentration of NO3-N/PO4-P ratio from 2000-2004; 2010-2013 at the Medlock (S0, S4 and S6).

S0 S4 S6 Period NO3-N PO4-P N/P NO3-N PO4-P N/P NO3-N PO4-P N/P S0 S0 S0 S4 S4 S4 S6 S6 S6 2000 1.99 0.14 14.21 3.75 1.02 3.68 3.34 0.77 4.34 2001 1.71 0.09 19.00 4.29 0.94 4.56 3.75 0.89 4.21 2002 1.80 0.08 22.50 5.23 1.02 5.13 3.94 0.58 6.79 2003 1.50 0.11 13.64 6.07 1.06 5.73 5.00 0.92 5.43 2004 1.64 0.13 12.62 5.46 1.02 5.35 3.89 0.65 5.98 2010 1.07 0.04 26.75 7.36 1.00 7.36 4.77 0.56 8.52 2011 1.24 0.06 20.67 6.49 0.84 7.73 4.78 0.57 8.39 2012 1.20 0.06 20.00 4.03 0.50 8.06 3.07 0.36 8.53 2013 1.14 0.05 22.80 5.03 0.76 6.62 4.19 0.60 6.98

Table 3-6: Average NO3-N/PO4-P ratio at Lord’s Brook, 2000-2006.

Mean N/P ratio Lord’s Brook 2000-2006

NO3-N PO4-P N/P ratio 2000 2.18 0.55 3.99 2001 2.84 0.52 5.44 2002 2.54 0.61 4.19 2003 1.86 0.42 4.43 2004 1.50 0.41 3.66 2005 2.44 0.79 3.11 2006 2.43 0.42 5.76

3.3.1.4 Summer vs Winter water chemistry

Average summer PO4-P and NO3-N concentrations showed that the nutrients (

Figure 3-6A & B) were mostly elevated in the summer especially for S4 and S6 as few winter concentrations were observed at S0. Higher summer nutrient concentration points to the effectiveness of the treatment works during the low discharge at summer months when dilution is reduced. Average concentrations of ammonia-N and suspended solids were highest during winter months which suggests that a rise in precipitation increases run-off plus releases from CSOs. There is no clear pattern with BOD (mgL¯¹) which suggests the influence of both point and diffuse pollution sources on the river. 65

A B

D E

Figure 3-6: Average winter vs summer water chemistry (A) PO4-P;(B) NO3-N; (C) Ammonia-N; (D)

BOD; (E) Suspended solids. Each icon, S0 (blue); S4 (red) and S6 (green) per site represents a year

3.3.2 Benthic macroinvertebrates

3.3.2.1 Invertebrate abundance and BMWP scores The abundance of benthic macroinvertebrates at S6 (Figure 3-7) between 2000 and 2008 showed the dominance of taxa in the following order: Oligochaeta >>

Baetidae > Chironomidae > Simulidae. The number of taxa identified at S6 and at

Lord’s Brook was 21 and 25 respectively. At Lord’s Brook (

Figure 3-8), a similar pattern of distribution to S6 was identified between 2000 and 2006. The dominance of tolerant benthic invertertebrates was indicated by the low BMWP scores and ASPT. Overall the BMWP score and ASPT placed the river at

S6 in the “polluted” category while Lord’s Brook was “moderately polluted”. This pattern agrees with the Environmental Quality Ratio (EQR) for the same period with

ASPT= 0.66, Number of taxa = 0.44 which placed the Medlock at the “moderate” pollution boundary for benthic macroinvertebrates (EA Data, 2016).

Figure 3-7: Benthic macroinvertebrate abundance at S6 between 2000 and 2008 and at Lord’s Brook between 2000 and 2006.

Figure 3-8: BMWP scores and ASPT at S6 and at Lord’s Brook

3.4 Summary i. On the basis of the WFD classification, the physico-chemical variables except

PO4-P were “good”. High PO4-P concentration placed water chemistry in the

“poor” category.

ii. The major change in the river was observed between S0 and S4 which are

respectively upstream and downstream of the WwTw. The treatment works

significantly increase PO4-P concentration at S4.

iii. Summer and winter results showed the effect of episodic pollution with a

deterioration in water quality during high winter precipitation.

iv. The benthic invertebrate community was indicative of moderate pollution as

indicated by the EQR and the BMWP scores.

3.5 Discussion The overall aim of this study was to assess the efficacy of water management measures on the river Medlock’s quality and ecology over time from Environment

Agency’s long term water quality dataset. This study encompassed more than a decade between 2000 and 2013 and allowed an assessment of the impact of regulatory policies specifically the European Water Framework Directive (WFD,

2000/60/EC).

Apart from NO3-N and PO4-P concentration which was higher at the downstream sites (S4 and S6) of the WwTW, other physico-chemical analysed in this study were similar at all sites. BOD was low throughout the study period indicating that sewage was effectively treated at the WwTW and there was little contribution from CSOs. Information obtained from the EA indicated significant efforts to reduce BOD concentrations in order to comply with the GQA (EA, personal communication 2016).and this has been achieved throughout the period of this study. Also, low BOD was shown by the high DO which in combination with the well-mixed water resulted in 80% saturation throughout the study period.

PO4-P concentration may have declined in the Medlock from 2000 to 2012, but would need to be justified with further analysis to ascertain the level of reduction.

However, the concentration of PO4 is still often higher than the 0.1PO4-PmgLˉ¹ WFD limit, especially in downstream urban areas of rivers such as the Medlock where this study showed concentration to average 0.5 mgLˉ¹. In common with many other rivers

(Howell, 2010), a major source phosphorus in the Medlock is from the (single)

WwTWs due to the lack of PO4-P removal from the effluent (Neal et al. 2008). James et al. (2012) showed that diffuse pollution is another reason for rivers in the Irwell 70

catchment failing to meet the legally required EU standards for phosphorus. As a result of continued PO4-P pollution from agriculture and runoff, most UK Rivers including the Medlock may not comply with the WFD until the next scheduled deadline of 2027 (Priestley 2015).

There are no statutory targets for NO3-N concentrations in UK surface waters under the WFD. However, the World Health Organization, (2007) states that NO3-N concentration in surface water is normally low at between 0 –18 mgLˉ¹ but can reach high levels as a result of agricultural runoff, refuse dump runoff or contamination with human or animal wastes. Within the EU, NO3-N concentrations in rivers declined by 0.8% each year over the period 1992 to 2012 following measures to reduce NO3-N from agricultural land and improvements in wastewater treatment

(European Environment Agency 2015).

Although NO3-N mirrors PO4-P concentration, a higher summer concentration as observed in the Medlock points to the effectiveness of the WwTW (Bowes et al.,

2015; Neal et al., 2005) which provides a constant effluent source and is less dilute in the summer low-flow months. A similar pattern to the Medlock was reported in the urban reach of the River Frome at Bristol (Bowes et al. 2009).

A summer-winter relationship showed that the concentration of ammonia- and suspended solids increased during winter in the Medlock. Increased concentration of ammonia could be linked to impact of CSOs (Mulliss et al. 1996;

Mullis et al. 1997) and high surface runoff (Martin, 1995) during the winter season

(Sigleo & Frick 2003) and Wang (2014) found high ammonia concentration during rainstorms in the Harlem river, New York. However, the increased concentration of ammonia is temporary as the River Medlock was highly oxygenated and hence conversion to NO3-N is rapid. The elevated suspended solids concentration may be also be linked to episodic storm events as was observed within the wider Irwell catchment (APEM, 2007). The higher suspended solids concentrations recorded in the

Medlock in 2000 and at 2012, which are above the EU FFD standards, may be linked

to the increased rainfall during these years as indicated by the Meteorological Office

(Online archive of the UK Meterological Office).

Biotic indices analysed for BMWP scores and ASPT from 2000 to 2008 at S6 indicate the river to be “impacted” by pollution and “moderately impacted” at Lord’s

Brook between 2000 and 2006 according to Hawkes (1997). The possible reasons for the low scores and degraded community could be linked to episodic discharge and urban runoff which transports suspended solids, PO4-P and other materials into the river (Paul & Meyer 2001; Walsh 2000). The benthic invertebrates were similar in the river and tributary and were dominated by pollution tolerant taxa including

Oligochaeta, Baetidae, Chironomidae and Simulidae. Goodnight, (1973) describes

Oligochaete as normal members of the stream biota and therefore, the classification of good and bad stream conditions will depend on their percentage contribution relative to the total stream biota. If the contribution of Oligochaete falls between 60% and 80%, this shows that the river was in bad condition and if < 60%, this would indicate good condition. On the basis of a high contribution of Oligochaete which constituted over 60% of the community in this study, the Medlock would fall under the bad category. Therefore, a subsequent chapter will explore the relationship between water quality and the degraded benthic invertebrate community.

3.6. Conclusion There had been no change in water quality in the River Medlock with time for the variables measured except for PO4-P. Although PO4-P declined, in particular at the most downstream (and hence urbanised) site, the concentration directly downstream of the treatment works exceeded WFD requirements. Thus, the

WwTWs was shown to be the major pollution source of PO4-P. As the concentrations of other physico-chemical variables were conducive to a diverse benthic invertebrate community, it is suggested that they were adversely affected by high precipitation which causes increased discharge, runoff and episodic

pollution from CSOs. These changing conditions are likely to destabilise the benthic

fauna and hence cause their impoverishment.

It is therefore apparent that the River Medlock has not improved in water quality

and does not comply with both the requirements of the WFD for good chemical and

ecological status. The EA data provided an overview of the river quality and

revealed high concentrations of phosphorus, the next chapter will investigate the

PO4-P dynamics in the river in an attempt to identify and quantify the major

sources of this contaminant.

Acknowledgements I am grateful for financial support from The National Open University of Nigeria.

Thanks to the Environment Agency, Warrington, UK for supplying the long-term

datasets used for this study.

3.7 References Bowes, M. J., Jarvie, H. P., Halliday, S. J., Skef, R. A., Wade, A. J., Loewenthal, M., … Palmer-felgate, E. J. (2015). Characterising phosphorus and nitrate inputs to a rural river using high-frequency concentration-flow relationships. Science of the Total Environment, 511, 608–620. http://doi.org/10.1016/j.scitotenv.2014.12.086

Bowes, M. J., Smith, J. T., Jarvie, H. P., Neal, C., & Barden, R. (2009). Changes in point and diffuse source phosphorus inputs to the River Frome (Dorset, UK) from 1966 to 2006. Science of the Total Environment, 407(6), 1954–1966. http://doi.org/10.1016/j.scitotenv.2008.11.026

Council of the European Union. (2000). Water Framework Directive 2000/60/EC - Official Journal of the European Communities.

European Environment Agency. (2015). Nutrients in Freshwater. European Environment Agency. European Environment Agency. Retrieved from http://www.eea.europa.eu/data-and-maps/indicators/nutrients-in- freshwater/nutrients-in-freshwater-assessment-published-6

Goodnight, C. (1973). The Use of Aquatic Macroinvertebrates as Indicators of Stream Pollution. Transactions of the American Microscopical Society, 92(1), 1–13.

Hawkes, H. (1997). Origin and development of the biological monitoring working party score system. Water Research, 32(3), 964–968. http://doi.org/10.1016/S0043- 1354(97)00275-3

Howell, J. A. (2010). The distribution of phosphorus in sediment and water downstream from a sewage treatment works. Bioscience Horizons, 3(2), 113–123. http://doi.org/10.1093/biohorizons/hzq015

James, P., Atkinson, S., Barlow, D., Bates, A., Comyn, F., Duddy, M., … Causer, K. (2012). The Irwell Catchment Pilot:The Rivers Return. Warrington, UK: The Environment Agency.

Jarvie, H. P., Whitton, B. A., & Neal, C. (1998). Nitrogen and phosphorus in east coast British rivers: Speciation, sources and biological significance. Science of the Total Environment. http://doi.org/10.1016/S0048-9697(98)00109-0

Mainstone, C., Parr, W., & Day, M. (2000). Phosphorus and River Ecology: Tackling Sewage Inputs. Prepared on behalf of English Nature and the Environment Agency.

Manchester City Council. (2007). Report for information. Manchester.

Manchester City Council. (2014). River Medlock Restoration.

Mullis, R., Revitt, D. M., & Shutes, R. B. E. (1997). The impacts of discharges from two combined sewer overflows on the water quality of an urban watercourse. Water Science & Technology, 36(8–9), 195–199. http://doi.org/10.1016/S0273- 1223(97)00599-4

Mulliss, R. M., Revitt, D. M., & Shutes, R. B. (1996). The impacts of urban discharges on the hydrology and water quality of an urban watercourse. Science of The Total Environment, 189–190, 385–390. http://doi.org/10.1016/0048-9697(96)05235- 7

Neal, C., Jarvie, H. P., Love, A., Neal, M., Wickham, H., & Harman, S. (2008). Water quality along a river continuum subject to point and diffuse sources. Journal of Hydrology, 350(3–4), 154–165. http://doi.org/10.1016/j.jhydrol.2007.10.034

Neal, C., Jarvie, H. P., Neal, M., Love, A. J., Hill, L., & Wickham, H. (2005). Water quality of treated sewage effluent in a rural area of the upper Thames Basin, southern England, and the impacts of such effluents on riverine phosphorus concentrations. Journal of Hydrology, 304, 103–117. http://doi.org/10.1016/j.jhydrol.2004.07.025

Paul, M. J., & Meyer, J. L. (2001). Streams in the urban landscape. Annual Review of Ecology and the Systematics, 32, 333–365.

Priestley, S. (2015). Water Framework Directive : achieving good status of water bodies (No. CBP 7246).

Sigleo, A., & Frick, W. (2003). Seasonal variations in river flow and nutrient concentrations in a northwestern USA watershed. … on Research in the Watersheds. US …, (Figure 1). Retrieved from 74

http://www.tucson.ars.ag.gov/icrw/Proceedings/Sigleo.pdf

Standard Committee of Analysts Publications, E. A. (2011). Index of methods for the examination of waters and associated Materials 1976-2011 Blue Book 236. Bristol: Environment Agency.

UKTAG Water Framework Directive. (2013). UK Technical Advisory Group on the Water Framework Directive Updated Recommendations on Environmental Standards River Basin Management (2015-21 ) Final Report. United Kingdom.

Walsh, C. J. (2000). Urban impacts on the ecology of receiving waters: a framework for assessment, conservation and restoration. Hydrobiologia, 431(2–3), 107–114. http://doi.org/10.1023/A:1004029715627

Wang, J. (2014). Combined Sewer Overflows (CSOs) Impact on Water Quality and Environmental Ecosystem in the Harlem River. Journal of Environmental Protection, 5, 1373–1389. http://doi.org/10.4236/jep.2014.513131

WFD-UK Technical Advisory Group (UKTAG). (2012). A revised approach to setting Water Framework Directive phosphorus standards.

Williams, A. E., Waterfall, R. J., White, K. N., & Hendry, K. (2010). Manchester Ship Canal and Salford Quays: industrial legacy abd ecological restoration. In L. C. Batty & K. . Hallberg (Eds.), Ecology of Industrial Pollution (pp. 276–308). Cambridge University Press.

World Health Organization. (2007). Nitrate and Nitrite in Drinking Water. Background Document for Development of WHO Guidelines for Drinking Water Quality, (WHO/SDE/WSH/07.01/16/Rev/1), 31. http://doi.org/10.1159/000225441

Wright, J. F., Moss, D., Armitage, P. D., & Furse, M. T. (1984). A preliminary classification of running water sites in Great Britain based on macroinvertebrate species and the prediction of community type using environmental data. Freshwater Biology, 14, 221–256. http://doi.org/10.1111/j.1365-2427.1984.tb00039.x

Bowes, M. J., Jarvie, H. P., Halliday, S. J., Skef, R. A., Wade, A. J., Loewenthal, M., … Palmer-felgate, E. J. (2015). Characterising phosphorus and nitrate inputs to a rural river using high-frequency concentration-flow relationships. Science of the Total Environment, 511, 608–620. http://doi.org/10.1016/j.scitotenv.2014.12.086

Council of the European Union. (2000). Water Framework Directive 2000/60/EC - 75

Official Journal of the European Communities.

Goodnight, C. (1973). The Use of Aquatic Macroinvertebrates as Indicators of Stream Pollution. Transactions of the American Microscopical Society, 92(1), 1–13.

Hawkes, H. (1997). Origin and development of the biological monitoring working party score system. Water Research, 32(3), 964–968. http://doi.org/10.1016/S0043- 1354(97)00275-3

Howell, J. A. (2010). The distribution of phosphorus in sediment and water downstream from a sewage treatment works. Bioscience Horizons, 3(2), 113–123. http://doi.org/10.1093/biohorizons/hzq015

James, P., Atkinson, S., Barlow, D., Bates, A., Comyn, F., Duddy, M., … Causer, K. (2012). The Irwell Catchment Pilot:The Rivers Return. Warrington, UK: The Environment Agency.

Mainstone, C., Parr, W., & Day, M. (2000). Phosphorus and River Ecology: Tackling Sewage Inputs. Prepared on behalf of English Nature and the Environment Agency.

Manchester City Council. (2007). Report for information. Manchester.

Manchester City Council. (2014). River Medlock Restoration.

Neal, C., Jarvie, H. P., Neal, M., Love, A. J., Hill, L., & Wickham, H. (2005). Water

quality of treated sewage effluent in a rural area of the upper Thames Basin, southern England, and the impacts of such effluents on riverine phosphorus concentrations. Journal of Hydrology, 304, 103–117. http://doi.org/10.1016/j.jhydrol.2004.07.025

Paul, M. J., & Meyer, J. L. (2001). Streams in the urban landscape. Annual Review of Ecology and the Systematics, 32, 333–365.

Priestley, S. (2015). Water Framework Directive : achieving good status of water bodies (No. CBP 7246).

Sigleo, A., & Frick, W. (2003). Seasonal variations in river flow and nutrient concentrations in a northwestern USA watershed. … on Research in the Watersheds. US …, (Figure 1). Retrieved from http://www.tucson.ars.ag.gov/icrw/Proceedings/Sigleo.pdf

Standard Committee of Analysts Publications, E. A. (2011). Index of methods for the examination of waters and associated Materials 1976-2011 Blue Book 236. Bristol: Environment Agency.

WFD-UK Technical Advisory Group (UKTAG). (2012). A revised approach to setting Water Framework Directive phosphorus standards.

Chapter 4 SOURCES OF PO4-P IN AN URBAN RIVER: COMBINED SEWER OVERFLOWS VS WASTEWATER TREATMENT WORKS

Abstract

This study examines concentrations and load of PO4-P in the urban River Medlock, Greater Manchester UK. This river has a history of pollution from the industrial revolution with significant contributions from wastewater treatment works (WwTW) and combined sewer overflows (CSOs). Data was obtained from the Environment Agency between 2000 and 2013, water samples were collected from the river every fortnight from March 2013 to April 2014 and a short term high resolution data was obtained at a single site over a period of four months. Concentrations of PO4-P in the river significantly reduced over the last decade but still show high PO4 concentrations >0.1mgLˉ¹ were observed, most commonly at all the sample sites below the WwTW where amounts reached 0.57 mgLˉ¹. PO4-P load varied from 1.01kgha⁻¹yr⁻¹ to 3.17kgha⁻¹yr⁻¹ and, on the basis of load estimate, about 90% came from the single WwTW discharge.

Keywords: PO4-P; River Medlock; PO4-P load; wastewater treatment works; Water Framework Directive.

4.1. Introduction

The adverse effects of elevated PO4-P has been a major environmental problem in water bodies for decades (Mainstone et al. 2000; Meybeck & Helmer 1989). These problems have been linked to diffuse pollution from urban runoff, storm water drains and combined sewer over flows (CSOs) in urban rivers by various authors including Barco et al. (2008); Hatt et al. (2004) and from discharge from wastewater treatment works (WwTWs) (Williams et al. 2010).

In the Northwest England, urban pollution affecting rivers was associated with the industrial revolution of the 19th century (Manchester City Council 2007;

Burton 2003). In the 21st century, these pollution effects were associated with point and diffuse sources in the river catchments (James et al. 2012). Following the implementation of various European Union Directives (Urban Wastewater

Treatment Directive, Freshwater Fisheries Directive and Nitrate Directive) aimed at reducing contaminants discharged to surface waterbodies, the concentration of trace metals; NO3-N, ammonia-N and biochemical oxygen demand have reduced.

However, levels of PO4-P have remained elevated as reported by the European

Union Water Framework Directive (WFD) requirements (Council of the European

Union 2000).

Efforts to reduce the PO4-P concentration to less than the WFD standard of

0.1mgLˉ¹ have largely failed (Bowes et al. 2009; Neal et al. 2005;Jarvie et al. 2006) due to contributions from WwTWs that lack tertiary treatment to remove this element from the effluent stream. While WwTW has been identified as a pollution source, the quantification of PO4-P from other sources including CSOs have yet to be determined. CSOs are storage drains which receive both urban runoff and wastewater and which are discharged into rivers during storms. Some of the effects of CSO discharges are seen in increased chemical pollution including PO4-P plus increased flow and discharge (Even et al. 2004; Wang 2014). In the UK, approximately 31,000 CSOs have been reported to discharge into fresh running watercourses (Marine Conservation Society 2011).

PO4-P concentrations from WwTWs (Bowes et al. 2005), other point sources

(Nyamangara et al. 2013; Hirsch 2012) and from agricultural land (Edwards &

Withers 2007; Edwards & Withers 2008) have been reported in the literature. There is however a significant research gap in quantifying the contribution to PO4-P load of

CSOs and WwTW given that CSOs are only monitored in the case of localised impacts on water quality and the biota (Marsalek et al. 2006). This study therefore investigates the dynamics of PO4-P released to the largely urbanised lower River

Medlock by assessing long term load over different periods and discharge regimes

using past data from the Environment Agency and primary data collection. The overall aim is to quantify the load of PO4 as P arising from the WwTW in order to assess its contribution to the river. The questions to be addressed in the study are as follows:

1. What is the relationship of PO4-P with discharge?

2. What is the spatial and temporal pattern of PO4-P concentration and flux?

3. What is the contribution to PO4-P load from, respectively, CSOs and the single

operational WwTW?

Study area: The River Medlock The River Medlock catchment (Figure 1-5) is heavily urbanised, in particular the lower reaches (CEH, 2014). The catchment area of the Medlock is 57.5 km2 and the average rate of flow in the river is 0.8m³s¯¹ (CEH, 2014). The River Medlock rises in the hills to the NE of Oldham in Greater Manchester (National Grid Reference

(NGR): SD 95308 05431). The Medlock flows for 22km through steep sided woodland and continues in a south westerly direction to discharge into the River Irwell immediately downstream of Manchester city centre (NGR: SJ 85781 97858). The study area encompasses both less and more heavily urbanised areas and extends from Mill

Brow Bridge, 6.6km from the river’s source to the confluence with the River Irwell

17.4 kilometres downstream (Figure 2-2). This reach was also chosen as it encompasses the single continuously operational wastewater treatment works at

Failsworth (NGR: SJ 89666 99802) plus twenty-nine CSOs (Personal Communication,

United Utilities, 2015). In the past, the river had a history of pollution especially from industrial effluent, and inadequately treated sewage (Douglas et al. 2002; MacKillop

2012;Williams et al. 2010). Rees & White (1993) also attributed pollution to localised storm events which led to frequent spill events from CSOs.

4.2. Methods

Phosphorus in the form of PO4-P was used throughout this study. This is the bioavailable form of P utilised by plants including algae (Jarvie et al. 2006).

The long term Environment Agency data were selected for three river sites within the study reach and the main tributary, Lord’s Brook. The EA data was complemented by bi-monthly sampling at six sites from March 2013 to April 2014 to encompass a full season. CSOs were identified during this study but water samples were not obtained due to safety considerations and no past measurements were available. High frequency spot sampling from a single site from August 2014 to

October 2014 in order to examine changes in PO4-P concentration during high and low discharge was carried out to examine the effect of rainfall events on PO4-P concentration and load.

4.2.1 Low resolution long term EA data Three sites designated S0, S4 and S6 were selected from a number of EA sampling locations on the Medlock as these were the only sites within the study area that bracket the WwTW. Site S0 is 3km above the WwTW, S4 and S6 are 0.8km and

5km downstream of the WwTW respectively. Lord’s Brook is 0.5km upstream of the

WwTW and was also assessed to determine its effect on the river although data was only available from 2000 to 2006. The sampling frequencies over the study period between 2000 and 2013 varied with each site. While S0 and S4 had datasets from 2000 to 2004 and from 2010 to 2013, there was no data from 2005 and 2009. S6 had complete data sets from 2000 to 2013.

Instantaneous continuous discharge records at 15 minute intervals was obtained from the EA at the single gauging station (NGR: SJ 849 975) on the Medlock to estimate PO4-P load at the three sites. Discharge readings for the three locations were determined on the basis of each site’s sub-catchment area as shown on Table

4-1. Each sub-catchment area was divided by the total catchment area (57.5km²) and

the result multiplied by the instantaneous discharge records obtained from the

gauging station. Daily discharge records from the WwTW were obtained from the

water company, United Utilities.

Table 4-1: Environment Agency sampling sites S0, S4 and S6, Lord’s Brook at the River Medlock

Station Name No. Catchment Catchment Distance Mean Mean area (km²) area as % of (km) from Discharge Discharge gauging source (m³s¯¹) (m³s¯¹) station 2000-2004 2010-2013 Lord’s Brook 2.58 4.5 12.07 0.03 0.03

Medlock Vale S0 37.20 65 12.23 0.50 0.46 Millstream Lane S4 43.90 76 13.04 0.59 0.55 Pin Mill Brow S6 54.40 95 17.40 0.73 0.68

4.2.1.1 Estimating load

PO4-P load was estimated using two methods (extrapolation and interpolation;

Littlewood 1992) in order to verify the outcomes of each. Estimates of PO4-P load to

the river were derived using the regression/rating curve method (Walling & Webb

1985). The extrapolation method was applied to the EA datasets due to the limited

number of PO4-P concentration measurements to provide estimates of the PO4-P

load. A relationship between the continuous daily discharge (m³s¯¹) and the PO4-P

concentration (mgL¯¹) was employed to generate continuous PO4-P flux (kgha ˉ¹

yearˉ¹) between the sites. The load was estimated using the following equation:

Load = K. Δt. Σ(Ci . Qi)

where K is a constant, Δt is the data time interval; Ci is the concentration of sample and Qi is the discharge at the time of sampling.

All concentration and discharge were transformed using the power law function

(extrapolation 1) and log-log regression (extrapolation 2) before the derivation of

rating curves. By using the rating curves, PO4-P concentrations were estimated for

every 15-minute interval which corresponded with the discharge records. In the

estimation of river loads, Walling & Webb (1985) assessed five interpolation methods

of nutrient load estimation (“Methods 1 to 5”) to determine their reliability. While

other methods were based on the estimates of time-weighted rather than on the flow- 82

weighted value, the “Method 5” interpolation technique was considered representative of conditions occurring between sampling occasions. The resultant load estimate will depend entirely upon the representativeness of the flow-weighted mean concentration value derived from a small number of samples. “Method 5” is also a preferred method recommended by the Paris Commission for assessing river inputs of Red List and other substances to the North Sea (Littlewood 1992).

Interpolation “Method 5” was also applied to the samples collected during this study- fortnightly and the high frequency samples using the following equation:

K = a conversion factor to account for (a) the period of load estimation and (b) units; time interval (in seconds) over which the load was calculated (kgyearˉ¹): (86400 = number of seconds in a day) x 1000 (correcting m³ to Litres))/106 (correcting mg to kg) = sample concentration; = discharge at sample time; = annual mean discharge for period of load estimate (record); n = number of samples; (i=1,…n) are the times at which the samples were taken. The resulting load estimates for both methods (extrapolation and interpolation) were compared to each other in order to examine differences or similarity in output.

4.2.2 Fortnightly spatial data

Water samples were obtained fortnightly from the River Medlock from March

2013 to April 2014 at the six monitoring locations as shown on Table 4-2.

Table 4-2: Description of sampling sites on the Medlock including values for average (no =23) of width, depth, velocity (V) and discharge (Q) of each site.

Station Name Station No. Catchment Dist. Ave. Average Average Average area (km²) (km) Width Depth V (ms-1) Q (m³s- from (m) (m) 1) Source Mill Brow S1 15.00 6.60 5.70 0.17 0.15 0.15 Park Bridge 8.20 0.25 0.11 0.23 Road S2 23.50 8.50 Daisy Nook 8.60 0.21 0.16 0.29 Garden S3 29.70 10.3 Millstream 8.80 0.27 0.18 0.43 Lane S4 43.90 13.00 Purslow Close S5 53.70 16.10 8.80 0.22 0.27 0.53 Pin Mill Brow S6 54.40 17.40 8.50 0.30 0.21 0.53 WwTW N.A N.A 12.60 N.A N.A N.A N.A

More sample sites were selected compared to the EA temporal data to obtain a higher spatial resolution and hence identify changes in flow patterns and PO4-P concentration from groups of CSOs and the WwTW. S1 to S3 are upstream and S4 to

S6 are below the WwTW. Two sites S4 and S6 correspond to the equivalent EA sites.

CSOs are monitored by the United Utilities’water Company and a simulated spill analysis (frequency, duration and volume of discharge) from CSOs was obtained from them. A summary of the results are presented in Table 4-3. The sub-catchment areas were determined using the Terrain Analysis System, GIS (Lindsay 2005) and a

50m (horizontal) and 0.1m (vertical) digital terrain model (DTM) (Centre for Ecology and Hydrology DTM 2015).

Table 4-3: Simulated prediction of frequency, duration and volume from CSOs (Source: United Utilities)

Sample sites No. of CSOs Frequency Duration Volume discharging into each (spills/yr) (hours) (m³) sample site

S1/S2 7 46 176.8 735,617 S3 8 285.5 598 346,912 S4 7 420.1 3547.9 585,608 S5 5 72.7 79.9 29,902 S6 1 1 0.6 1,005

4.2.3 Data collection For the fortnightly samples, one litre water sample was collected from the six sample sites in acid-washed (10% HCl for 24 hours) polypropylene bottles. Part of the water samples was filtered through 0.45µm Millipore cellulose acetate filter for the measurement of PO4-P and 40ml unfiltered sample was preserved for the analysis of total phosphorus (TP) by acid digestion (Mackereth et al. 1978). Total phosphorus was analysed for the estimation of TP load in the river.

Water velocity was measured using the float method i.e. by recording the time taken for the float to travel over a given distance (10m) along the river. Fortnightly discharge was calculated for each sub-catchment area on the basis of their relationship to the total catchment area. Discharge records used for the sub- catchment area obtained at the continuously gauged Environment Agency site was considered a viable option because the discharge calculated based on the river measurement varied extensively due to poor access, and at periods of high flows.

PO4-P samples were processed within 24 hours of the sample collection and analysed using a SEAL Auto Analyzer 3 High Resolution instrument (SEAL

Analytical Ltd, Southampton). Detection limits for PO4-P measured as P is 0.004 mgLˉ¹. For further information on the methods employed by the auto analyser see

SEAL Analytical (2013).

4.2.4 High resolution temporal dynamics

Hydrograph monitoring and precipitation: A 15-minute duration discharge record was obtained from stage records supplied by the EA’s gauging station. Precipitation data over the 15-minute periods was obtained from the Whitworth Meteorological

Observatory, Manchester.

Spot sampling: Water samples were collected from the EA gauging station over a range of low and high flow conditions between August 2014 and October 2014. There was no direct access into the river; therefore, the water sample was obtained by lowering a bucket into the river immediately upstream of the gauging station. These samples were then decanted into a one-litre sample container for subsequent analysis. The sample was analysed as described above.

4.2.5 Data analysis Statistical analyses were carried out using Microsoft Excel 2013, SPSS (IBM, 2013) or

GraphPad Prism version 6.

4.3 Results The results are presented in four sections- low resolution long term EA data for the river and the tributary (between 2000 and 2013), fortnightly sampling (2013-

2014), high frequency three month sampling (2014, August to October) and a comparison of the results. The rationale is that the long-term changes in the river

(section 4.3.1), the seasonal pattern, concentration-discharge relationship, load at different locations provides an overview of the river’s conditions with time; the fortnightly data (section 4.3.2) provides changes over an annual cycle with more spatial resolution; high frequency spot sampling was determined so as to understand temporal changes in the river at the gauged site in relation to rainfall and discharge

(section 4.3.3). The last section 4.3.4 compares the PO4-P load on the Medlock for the entire period of study between 2000 and 2014.

4.3.1 Low resolution long term data

The mean PO4-P concentration between 2000 and 2004, and from 2010 to 2013 was shown on Figure 3-2H (Chapter 3). For both periods, the concentration of PO4-P increased with site S6>S4>>S0 and only S0 complied with the WFD requirements of

<0.1 mgLˉ¹. A one-way ANOVA to distinguish the concentration between the three study sites revealed a highly significant (p < .01) difference in the period 2000 to

2004 (F2,12 = 111.27) and from 2010 to 2013 (F2,9 = 28.97). At site S6, where continuous data was available from 2000 to 2013, Pearson correlation showed a significant (p <

.05, r = -0.72) reduction in the concentration of PO4-P between 2000 and 2013.

Average summer and average winter PO4-P concentration at the three sites (

Figure 3-6A (Chapter 3) shows that the concentration was elevated in the summer at all sites, with some periods of elevated winter concentrations at S0.

Elevated concentrations suggest the impact of discharge from the WwTW during low flow summer periods which is indicative of continuous sewage discharge.

4.3.1.1 PO4-P concentration-discharge relationship

The 14-year measurement of PO4-P concentration and discharge dataset was divided into two-five-year and one-four-year period (from 2000 to 2004, 2005 to 2009 and 2010 to 2013). This duration was considered appropriate as the UK water industry operates on five-yearly cycles called Asset Management Plans (AMP) as directed by the Office of Water Services (OFWAT) (http://www.ofwat.gov.uk/). The

AMP is a plan which delivers specific objectives on water infrastructure by the combination of multi-disciplinary management techniques over the water life cycle by the water companies (dream report.net, 2014). In this study the AMP covers AMP

3 (2000-2004); AMP 4 (2005-2009) and AMP 5 (2010 -2014). Each of the five-year data subsets was modelled as power law functions for the three sites S0, S4 and S6 (Figure

4-1, Figure 4-2, Figure 4-3) (Bowes et al., 2009). The line of best fit shown by the coefficient of determination R² for the concentration-discharge plots are described.

A weak positive correlation occurred at S0 (Figure 4-1) from 2000-2004 (r =

0.46, n=56, p=0.0004) and from 2010 to 2013 (r = 0.59, n=44, p=0.0001) which indicate diffuse source pollution such as agricultural runoff upstream. However, a negative correlation between PO4-P concentration and discharge occurred at S4 (2000-2004, r= -0.65, n=54, p=0.000 & from 2010 to 2013, r = -0.57, n=46, p = 0.000) (Figure 4-2) and at S6 (2000-2004, r= -0.40, n=58, p=0.0017; from 2005-2009; r= - 0.47, n=60, p=0.0001 & from 2010 to 2013 r= -0.52, n=49, p=0.0001) (Figure 4-3). The results showed the dilution behaviour of the river as both downstream sites (S4 and S6) showed a decrease in PO4-P concentration with increasing discharge and suggest pollution from the WwTW situated above these sites.

Figure 4-1: PO4-P concentration versus discharge at S0, 2000-2004 and at 2010-2013 estimated by extrapolation method using power law functions.

Figure 4-2: PO4-P concentration versus discharge at S4, 2000-2004 and at 2010 -2013 estimated by extrapolation method using power law functions.

Figure 4-3: PO4-P concentration versus discharge at S6, 2000-2004, 2005-2009 and 2010-2013 estimated by extrapolation method using power law functions.

4.3.1.2 PO4-P load for the low resolution dataset

Figure 4-4 shows the estimated PO4-P load into the River Medlock at the three sample sites by using the methods of extrapolation and interpolation. The two methods were found to be similar in output as no significant (p > 0.05) difference was found. This suggests that either method was a suitable predictive tool for the estimation of load.

Table 4-4 presents the mean and range of PO4-P load for the three EA study sites, the WwTW, and the main tributary on the study location, Lord’s Brook, by interpolation “Method 5”. The mean rate of PO4-P load input at each site was computed using extrapolated method (power law function) at all the sites over the duration of monitoring and by dividing each catchment’s export rate by the annual discharge. The results for the low resolution data between 2000 and 2004, S0 to S6,

Lord’s Brook had an estimated load range of 0.60 to 3.13 kghaˉ¹yrˉ¹ and between 2010 and 2013; a range of 0.26 to 2.16 kghaˉ¹yrˉ¹. Between 2010 and 2013, the increased load at S4 may be attributed to increased release of PO4-P from the WwTWs.

Figure 4-4: Comparison of Interpolation Method 5 and two extrapolation methods to estimate PO4-P load kg haˉ¹ yrˉ¹ at sites S0, S4 and S6.

Table 4-4: Mean and range of PO4-P load estimated at sites S0, S4, S6, at the WwTW and Lord’s Brook for 2000 to 2004 and 2010 to 2013.

PO4-P kghaˉ¹yrˉ¹ Sites 2000-2004 2010-2013

mean range mean range Lord’s 1.91 0.86-2.80 Brook WwTW (2012=1.21) 1.18-1.21 (2013=1.18) S0 0.6 0.35-1.04 0.26 0.18-0.37 S4 3.13 2.63-3.83 2.16 1.55-3.18 S6 2.52 1.83-3.17 1.45 1.01-1.95

4.3.2 Fortnightly spatial data

The concentration of PO4-P (mgLˉ¹) at the six sample locations S1-S6 is shown on Figure 4-5. Higher concentrations of PO4-P were obtained at sites S4, S5 and S6 which were located downstream of the WwTWs compared to sites S1, S2 and S3 which were upstream. The difference was confirmed in a one-way analysis of variance (ANOVA) which showed a highly significant (p <0.001) difference between the upstream and downstream sites sites (F5, 132 =13.2). Post hoc Tukey multiple comparisons tests showed a high significant difference (p<0.0005) between

S1 and S4 (p=0.0004) and between S1 and S6 (p=0.0008). Differences were found between S2, S3 and the downstream sites, p= 0.005).

1 .8

1 .6

1 .4

)

1 -

L 1 .2 g

m 1 .0

(

0 .8

- 4

O 0 .6 P 0 .4

0 .2

0 .0 S 1 S 2 S 3 S 4 S 5 S 6

S a m p le L o c a tio n s

Figure 4-5: PO4-P concentration at sample locations S1 to S6 in relation to distance from the source of the River Medlock. Box and whiskers represent 25% and 75%, median, minimum and maximum values.

The relationship between PO4-P concentration and discharge at sites S1 to S6 is shown on Figure 4-6. While S1, S2 and S3 showed no (p > .05) correlation, S4, S5 and

S6 showed a significant (p < .05) negative correlation (S4 r= -0.57; S5 r= -0.48; S6 r = -

0.56) of PO4-P with discharge. This pattern is similar to data obtained from the EA and shown on 0. This decrease in concentration of PO4-P with increased discharge is a further suggestion of the influence of the discharge from the WwTW.

Figure 4-6: PO4-P concentration vs discharge at sites S1 to S6. Regression line showed a decline in concentration with increased discharge.

The summer and winter relationship (Figure 4-7) showed higher PO4-P concentration during the summer compared to the winter season. During the low rainfall summer period there is limited dilution of PO4-P whereas during winter, higher concentration suggests overflows from combined sewers especially during storms and of runoff from urban surfaces.

Figure 4-7: Summer and winter PO4-P concentration (mgLˉ¹; note log scale) in the River Medlock using the fortnightly datasets collected between March 2013 and April 2014.

4.3.2.1 PO4-P load for fortnightly spatial data at sample sites

Table 4-5 shows the PO4-P load (kgdayˉ¹) contributions from the upstream site and WwTW. The concentration and discharge data used for the calculation of the latter was obtained from the EA and United Utilities. From the table, the total PO4-P load entering the river downstream of the treatment works is between 14.54 and

29.48 kgdayˉ¹ (average 22 kgdayˉ¹). Therefore percentage contribution from the

WwTW is calculated to be between 71 and 99% and shows the treatment works as the major contributor of PO4-P to the Medlock at site S4

Table 4-5: Contribution of PO4-P load from the WwTW.

PO4-P Load kgdayˉ¹

Discharge Total load % WwTW Date S3 WwTW (m³sˉ¹) (S3+STW) Contribution

09-04-13 0.13 8.66 20.82 29.48 70.62 24-04-13 0.12 0.25 19.96 20.21 98.77 26-06-13 0.12 0.02 14.52 14.54 99.85 13-08-13 0.25 1 15.34 16.35 93.88 25-11-13 0.17 0.68 17.73 18.41 96.29 11-12-13 0.26 0.61 25.57 26.18 97.68 14-01-14 0.24 5.7 22.12 27.82 79.51 18-02-14 0.9 1.14 17.87 19.01 94.00 Min 14.54 70.62 Max 29.48 99.85 Average 21.50 91.325

4.3.3 Temporal dynamics

PO4-P concentration vs discharge relationship (

Figure 4-8) showed a significant negative correlation (r= -0.269, p <0.05) which usually indicates dilution from a point source particularly the WwTW (Bowes et al.

2015).

WFD standard 0.1mgL¯¹.

Figure 4-8: Concentration of PO4-P vs discharge measured from 1st August and 31st October 2014. 96

4.3.3.1 Hydrograph of chemical concentration vs discharge

In the hydrograph of Figure 4-9 based on the 15-minute continuous discharge record was plotted against spot sample analysis of PO4-P collected over the same period. The hydrograph/PO4-P plot revealed higher concentration of PO4-P especially at low discharge levels which could indicate WwTW operation. However, higher discharge records with corresponding higher PO4-P concentration was recorded on the 10th August 2014, concentration of PO4-P (0.73mgL¯¹) with increased discharge

(6.82m³s¯¹) which suggests contributions from CSOs following increased volume arising from rainfall events. The highest PO4-P concentration in the river during the study period was recorded at 1.17mgL¯¹ with a discharge of 0.29m³s-¹ (5th August

2014). This value is 12-fold higher than the WFD standard of 0.1mgL⁻¹.

Figure 4-9: Hydrograph of 15-minute discharge at the gauging station and spot sample analysis of

PO4-P measured over the study period from 1st August 2014 to 31st October 2014. Sharp discharge spikes and high concentration suggest diffuse pollution sources.

4.3.3.2 Rainfall and discharge

The total precipitation recorded from 1st August to 31st October was

209.75mm with a mean 0.02mm, with minimum and maximum rainfall ranging from no precipitation to 7.37mm during the sample duration.

The hyetograph (Figure 4-10) is shown of precipitation over the sampling period. A mean discharge of 0.49m3s¯¹ was recorded throughout the sampling period with a minimum and maximum discharge of 0.16m3s¯¹ and 6.91m3s¯¹ respectively.

During these times, 69 peak discharges between 0.16m3s¯¹ and 6.91m3s¯¹ out of 8825 measurements were recorded manually. A significant correlation was found between discharge and precipitation (Pearson correlation, n=8825, p < .01) which, as expected, indicates that precipitation influenced discharge in the river.

Figure 4-10: Hyetograph of 15-minute precipitation and discharge over the study period from 1st August to 31st October 2014.

4.3.4 Comparing PO4-P load and concentration

Table 4-6 compares the PO4-P load over the duration of study from EA datasets, bimonthly sampling and the short-term (between 1st August and 31st

October 2014) spot sampling at the gauging station. S6 has been estimated based on the value obtained from the gauged station. The mean PO4-P load estimated at the river between 2000 and 2004 ranged from 3.13 to 0.6 kgPhaˉ¹yrˉ¹; between 2010 and

2013, load ranged from 2.16 to 0.26 kgPhaˉ¹yrˉ¹; and, in 2014, the range was from 0.92 to 0.26kgPhaˉ¹yrˉ¹.

Table 4-6: Summary of mean load/range of PO4-P for long term EA data (2000-2004, 2010-2013), fortnightly data (2014 for sites S4 and S6) and 2014* (high resolution spot sampling expressed kgPha⁻¹yr⁻¹ for comparison).

kgPha⁻¹yr⁻¹

Sample locations - Catchment area -EA & fortnightly as % of EA sampling Gauging Station 2000-2004 2010-2013 2014 2014* WwTW 1.20 0.94 S0 65 0.60 0.26 0.82 S4 76 3.13 2.16 0.92 0.96 S6 95 2.52 1.45 0.68 1.2

The results show that between 2000 and 2014, the load of PO4-P in the river decreased with time across the sites. However, between sites, S4 was significantly higher than other sites. Between S4 and S0 and from 2000 to 2004, the load at S4 was five times higher (3.13 kgPha⁻¹yr⁻¹) than the load measured at S0, having a mean load of 0.60 kgPha⁻¹yr⁻¹ and, eight times much more between 2010 and 2013. For S4 and S6 and between 2000 and 2004, S4 was 19% higher than S6 and by 33% between 2010 and 2013. In 2014, the difference between both sites was by 26%. These values further emphasized the relative impact of the WwTW discharge on the river, especially at site S4 which is directly below the WwTW. However, an estimated load value from high frequency sampling in 2014 showed the load at S4 was lower than what was obtained at S6, suggesting the influence of CSOs or other sources during episodic

(rainfall) events contributing more PO4-P between S4 and S6.

The change in concentration over two decades is presented in Table 4-7. The % decrease in the concentration of PO4-P (mgLˉ¹) was calculated over a period of three decades in the River Medlock based on mean PO4-P concentrations for EA sites and the WwTW. Between two decades, i.e. 2002 & 2012, 2003 & 2013 (Figure 4-11), more than a 60% reduction in PO4-P concentration was observed in the effluent discharged from the WwTW. These changes were also reflected in the PO4-P concentration recorded in the river with > 30% reduction at S4 and S6.

With reference to simulated CSO spill analysis obtained from the water company (Table 4-3), spatial analysis of PO4-P concentration in this study showed that CSO discharge volume and duration do not significantly increase the concentration of PO4-P. Lower PO4-P concentrations was analysed for the upper S1,

S2 and S3 and higher concentrations were recorded downstream of the WwTW at S4,

S5 and S6. The impact of CSOs is greatest during short-duration episodic conditions.

During the high frequency sampling regime, high PO4-P was captured at one of the periods.

Table 4-7: Summary of mean PO4-P concentration (mgLˉ¹) measured at the WwTW and other EA sites between 2000 and 2012, 2003 and 2013, 2004 and 2014.

PO4-PmgLˉ¹/Period

Sites 2002 2012 2003 2013 2004 2014 WwTW 3.14 1.21 3.89 1.18 S4 1.02 0.5 1.16 0.76 1.02 0.57 S6 0.58 0.36 0.92 0.6 0.65 0.45

% change in PO4-PmgLˉ¹

(2002/2012) (2003/2013) (2004/2014)

WwTW -61.46 -69.67 S4 -50.98 -34.48 -44.12 S6 -37.93 -34.78 -30.77

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Figure 4-11: Mean PO4-P concentration in the River Medlock between 2002 and 2012, 2003 and 2013 at S4, S6 and at the WwTW

4.4 Discussion

High PO4-P concentration at the Medlock has been a major problem with regard to the river’s compliance with the WFD. The overall aim of this study was to determine the load of PO4-P entering the river from CSOs and the single operational

WwTW, and to examine the effects of seasonality. A further aim was to confirm the changes in PO4-P concentration and load in the river.

The long term and fortnightly measurements suggest that the sites below the

WwTWs had higher concentration of PO4-P (> 0.1mgL¯¹) and the main tributary

Lord’s Brook contributed a very small amount of PO4-P to the river.

Although long term datasets obtained from the EA were incomplete for some years at S0 and S4, analysis at the lowermost site S6, indicated a decrease in the concentration of PO4-P with time. Over two decades more than a 60% reduction in

PO4-P concentrations was shown to have occurred at S6 and this is highly likely to have resulted from a decrease in releases from the WwTW as there are no other major sources of PO4-P. Between these periods, more than 30% reduction were recorded at the site S4 which is 0.5km below the treatment works. Information provided by the

EA indicated there was no limit for PO4-P prior to year 2000 and therefore no 101

improvement plan was initiated for its discharge to rivers. However, stringent permit requirements were enforced under the UK River Ecosystem classification for the reduction of ammonia from 6mgL¯¹ to 3 mgL¯¹, BOD, which reduced from 30mgL¯¹ to

15mgL¯¹ and suspended solids which reduced from 45mgL¯¹ to 35 mgL¯¹. The EA suggests that enforcement of these standards could have resulted in a reduction in

PO4-P (EA, personal communication 2016).

Various suggestions to reduce PO4-P in UK rivers have been including the

United Utilities (DEFRA 2014; DEFRA 2012). The construction of underground storage tanks which will hold excess water and improve water quality entering the

Manchester ship canal is proposed by United Utilities for the Salford area of the city.

New or improved technologies such as pile cloth media filtration; membrane filtration; ballasted coagulation; nano-particle embedded ion exchange; immobilised algal bioreactor; and absorption media reed beds aimed at reducing phosphorus from

WwTWs are currently being evaluated at some Universities in the UK (WWT, March

2016). Tertiary treatment has been shown to reduce concentration of PO4-P elsewhere such as in the River Kennet and the River Thames (Jarvie et al., 2002) and could be applied to the WwTWs on the Medlock. As at 2010, the River Medlock fell within

44% of the rivers in North West England that have average PO4-P concentrations

>0.1mgLˉ¹ (Rothwell et al. 2010) which indicates that PO4-P contamination is a common problem in the region’s rivers, in part due to the lack of tertiary treatment of effluent from the WwTWs.

The concentration-discharge relationship supports the fact that the WwTW was the major point source. Howell (2010) and Halliday et al.(2014) pointed out that monitoring sites which directly received discharge from WwTW contained the highest PO4 concentration. Fortnightly sampling over a single season showed no relationship between discharge and PO4-P concentration at the upstream sites.

However, the long term EA datasets indicate that PO4-P sometimes increased with increasing discharge (Figure 4-1, 2010-2013). This suggests the influence of diffuse source impacts (Bowes et al. 2009). While concentrations can be reduced from point

102

sources such as WwTWs, diffuse sources from agriculture and from urban areas at the Medlock (Figure 1-5) runoff are more difficult to control as they would require changes to the use of PO4-P, including the use of fertilizers by the agriculture industry and other domestic applications of PO4-P.

The EA average summer and average winter PO4-P concentrations revealed that the highest concentrations occurred during the summer months when the river flow was lowest. The winter fortnightly datasets may suggest that during storm events the concentration of PO4-P could increase in the river due to remobilisation of the sediments (Bowes et al. 2015; Bowes et al. 2008).

PO4-P load measured during the high resolution sampling showed a higher load compared to those monitored during the long term and fortnightly sampling periods. This may be attributed to the remobilisation of particulate bound phosphorus from the bed sediment during flushing events where, large amounts of particulates were re-suspended (Brunet & Astin, 1998). Thus, ecological degradation occurs from the interaction between urban surfaces and other instream processes (Mulliss et al. 1996). The study period for the high resolution sampling

(August-October, 2014) was reported to be the wettest months in 2014 by the UK

Meteorological Station. Since August is the season of planting in the UK (Farming and Countryside Education (FACE) 2007), fertiliser application prior to high rainfall could increase nutrient load. Hence, timing of application is crucial to nutrient export (Beaulac & Reckhow 1982). Such diffuse pollutants pose a particular problem as they are generally widespread, hard to detect and to quantify (Beven et al. 2005). This study may suggest the influence of diffuse pollution, probably from headwaters, as well as from CSOs (based on data obtained from the United

Utilities’ company) during high flow conditions.

CSOs vs WwTW Due to the risk of sampling directly from CSOs especially during discharge, the contribution of PO4-P from the WwTW was determined. This method examined the importance of PO4-P load from upstream locations and, export from the WwTW 103

into the river. The results indicate that the WwTW exported higher PO4-P (average of 92%) load compared to other sources in particular CSOs, although storm drains may make a small contribution (Houston et al. 2011; Walsh et al. 2001). In this study, the shorter high frequency sampling between August and October 2014 indicate the influence of CSOs in providing very high concentration of PO4-P especially during storm conditions when CSOs are operational. This further suggests that, CSOs operated mainly during storm conditions. Therefore short duration sampling regimes can be used to assess the impact of CSOs under differing flows rather than a routine sample regime which may miss such events.

Comparison of phosphorus load in Medlock with other rivers The load of total phosphorus in the River Medlock measured during the fortnight sampling was compared with other studies of urban, semi-urban and agricultural catchments (Table 4-8). Table 4-8 indicates that semi-urban sections of the

River Medlock (S1/S2) were comparatively low compared to some other rivers such as Pevensey Levels which was ten times higher, and twice higher for River Ant.

Urban sections of the river Medlock (S3-S6) had lower TP load compared to other urban rivers with a ratio of 0.65:3. TP load recorded from urban construction activities and from agriculture were shown to be high while woodland had the lowest export rate (Table 4-8) on the basis of low fertiliser application and high infiltration rates (Line et al. 2015; Johnes 1996). Approximately 80% of total phosphorus recorded at River Avon (Bowes, Hilton, Irons, & Hornby, 2005) and

Pevensey Levels were associated with WwTWs.

This study showed the relative importance of point source pollution sources to the Medlock. However, several challenges were encountered especially when information on CSOs (i.e. CSO location, concentration of discharge, spill volumes and frequencies) were required from both the EA and the water companies. A similar pattern faced during this study has been encountered by other workers and presented as a policy document (Marine Conservation Society 2011) which calls for

104

the cooperation of the regulators, and the water company in order to work effectively to facilitate improvements in water quality.

With the provision of available datasets, future work could include the use of

Geographical Information Systems (GIS). The concentration of PO4-P at the various locations could be superimposed on a GIS map, together with the CSO locations and other information obtained on the river infrastructure. This information would represent high risk conditions within the shorted possible time. It is also important to determine the relative contribution of PO4-P concentration present in runoff and the river.

105

Table 4-8: Published phosphorus load based on different land uses.

Range Mean Land use Catchment (TP Source (TP kghaˉ¹yrˉ¹) (km²) kghaˉ¹yrˉ¹) Urban 0.56-3.36 (Donigian et al., 1994) (Commercial) Urban (Residential) 1.30 (EPA, 1983) Urban (Residential) 0.025 2.30 (Line et al., 2015)

Urban (Residential) 0.96 (Hartigan et al. 1983)

Urban 3.40 (EPA, 1983) (Commercial) Urban (Golf courses & 3.00 (Line et al., 2015) construction sites) 0.083

Semi-urban 46.90 (Bales et al. 1999) 0.69 Agriculture 0.10-3.25 0.94 (Beaulac & Reckhow 1982) (Cropping) Agriculture (Cropping & 0.14 (Ierodiaconou et al., 2005) irrigated pasture ) Agriculture <0.01-4.90 0.82 (Beaulac & Reckhow, 1982) (Pasture) Agriculture 4.30 (Line et al., 2015) (Pasture ) 0.062 Woodland 0.02 (Johnes, 1996) Woodland 0.033 1.0 (Line et al., 2015) Slurry and farm 31.90 (Tamminga, 1992) yard manure Semi-urban (Pevensey Levels) 56 6.10 (Mainstone et al., 2000) Semi-urban (River Ant) 49.3 1.00 (Johnes et al. 1994)

S1 (semi-urban) 0.46 S2 (semi-urban) 0.40 S3 (urban) 0.23 This study S4 (urban) 2.20 S5 (urban) 1.92 S6 (urban) 57.5 1.55

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4.5 Conclusion

This study has shown the concentration and load of PO4-P decreased with time in River Medlock. However, the river is yet to comply with the WFD requirements for PO4-P. The WwTW is the largest source of PO4-P load in the river and its contribution is greatest during the summer due to reduced discharge.

Evidence was also presented of significant PO4 contribution from CSOs during episodic storm conditions. High precipitation during fertiliser application period could also increase diffuse pollution. The comparison of total PO4-P load from the

Medlock with other areas showed that the Medlock was influenced by a combination of pollution sources enhanced by modified urban areas. The most cost-effective reduction in PO4-P concentration in the River Medlock could be achieved by investing in technologies which remove PO4-P from the treatment works effluent and also of reducing discharge entering the river. One option could be the use of wetlands to reduce the concentration of PO4-P in some parts of the river where land is available such as Clayton Vale.

Acknowledgements I am grateful for financial support from The National Open University of Nigeria.

Furthermore, thanks to the Environment Agency for the long term data sets,

Warrington and water company, United Utilities for supplying information on the

CSOs.

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Beven, K., Heathwaite, L., Haygarth, P., Walling, D., Brazier, R., & Withers, P. (2005). On the concept of delivery of sediment and nutrients to stream channels. Hydrological Processes, 19(2), 551–556. http://doi.org/10.1002/hyp.5796

Bowes, M. J., Hilton, J., Irons, G. P., & Hornby, D. D. (2005). The relative contribution of sewage and diffuse phosphorus sources in the River Avon catchment, southern England: Implications for nutrient management. Science of the Total Environment, 344(1–3 Spec. Iss), 67–81. http://doi.org/10.1016/j.scitotenv.2005.02.006

Bowes, M. J., Smith, J. T., Jarvie, H. P., & Neal, C. (2008). Modelling of phosphorus inputs to rivers from diffuse and point sources. Science of the Total Environment, 395, 125–138. http://doi.org/10.1016/j.scitotenv.2008.01.054

Brunet, R. C., & Astin, K. B. (1998). Variation in phosphorus flux during a hydrological season: The River Adour. Water Research, 32(3), 547–558. http://doi.org/10.1016/S0043-1354(97)00317-5

Burton, L. R. (2003). The Mersey Basin: An historical assessment of water quality from an anecdotal perspective. Science of the Total Environment, 314–316(3), 53– 66. http://doi.org/10.1016/S0048-9697(03)00094-9

Council of the European Union. (2000). Water Framework Directive 2000/60/EC - Official Journal of the European Communities.

DEFRA. (2012). Waste water treatment in the United Kingdom. Defra. United Kingdom.

DEFRA. (2014). Water Framework Directive implementation in England and Wales: new and updated standards to protect the water environment. UK. Retrieved from www.gov.uk/government/publications

Douglas, I., Hodgson, R., & Lawson, N. (2002). Industry, environment and health through 200 years in Manchester. Ecological Economics, 41(2), 235–255. http://doi.org/10.1016/S0921-8009(02)00029-0

Edwards, A. C., & Withers, P. J. A. (2007). Linking phosphorus sources to impacts in different types of water body. Soil Use and Management, 23(SUPPL. 1), 133–143. http://doi.org/10.1111/j.1475-2743.2007.00110.x 108

Edwards, A. C., & Withers, P. J. A. (2008). Transport and delivery of suspended solids, nitrogen and phosphorus from various sources to freshwaters in the UK. Journal of Hydrology, 350(3–4), 144–153. http://doi.org/10.1016/j.jhydrol.2007.10.053

Even, S., Poulin, M., Mouchel, J. M., Seidl, M., & Servais, P. (2004). Modelling oxygen deficits in the Seine River downstream of combined sewer overflows. Ecological Modelling, 173(2–3), 177–196. http://doi.org/10.1016/j.ecolmodel.2003.08.019

Farming and Countryside Education (FACE). (2007). Discovering Farming in Britain. United Kingdom.

Halliday, S. J., Skeffington, R. A., Bowes, M. J., Gozzard, E., Newman, J. R., Loewenthal, M., … Wade, A. J. (2014). The water quality of the River Enborne, UK: Observations from high-frequency monitoring in a rural, lowland river system. Water (Switzerland), 6(1), 150–180.

Hartigan, J. P., Quasebarth, T. F., & Southerland, E. (1983). Calibration of NPS Model Loading Factors. Journal of Environmental Engineering, 109(6), 1259–1272. http://doi.org/10.1061/(ASCE)0733-9372(1983)109:6(1259)

Hatt, B. E., Fletcher, T. D., Walsh, C. J., & Taylor, S. L. (2004). The influence of urban density and drainage infrastructure on the concentrations and loads of pollutants in small streams. Environmental Management, 34(1), 112–24. http://doi.org/10.1007/s00267-004-0221-8

Hirsch, R. M. (2012). Flux of Nitrogen, Phosphorus, and Suspended Sediment from the Susquehanna River Basin to the Chesapeake Bay during Tropical Storm Lee, September 2011, as an indicator of the effects of reservoir sedimentation on water quality. Retrieved from http://pubs.usgs.gov/sir/2012/5185/

Houston, D., Werritty, A., Bassett, D., Geddes, A., Hoolachan, A., & McMillan, M. (2011). Pluvial (rain-related) flooding in urban areas: the invisible hazard. Retrieved from http://www.jrf.org.uk/sites/files/jrf/urban-flood-risk-full.pdf

Howell, J. A. (2010). The distribution of phosphorus in sediment and water downstream from a sewage treatment works. Bioscience Horizons, 3(2), 113–123. http://doi.org/10.1093/biohorizons/hzq015

James, P., Atkinson, S., Barlow, D., Bates, A., Comyn, F., Duddy, M., … Causer, K. (2012). The Irwell Catchment Pilot:The Rivers Return. Warrington, UK: The Environment Agency.

Jarvie, H. P., Neal, C., Williams, R. J., Neal, M., Wickham, H. D., Hill, L. K., … White, J. (2002). Phosphorus sources, speciation and dynamics in the lowland eutrophic River Kennet, UK. Science of the Total Environment, 282–283, 175–203.

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Jarvie, H. P., Neal, C., & Withers, P. J. A. (2006a). Sewage-effluent phosphorus: A greater risk to river eutrophication than agricultural phosphorus? Science of the Total Environment, 360(1–3), 246–253. http://doi.org/10.1016/j.scitotenv.2005.08.038

Jarvie, H. P., Neal, C., & Withers, P. J. A. (2006b). Sewage-effluent phosphorus: A greater risk to river eutrophication than agricultural phosphorus? Science of the Total Environment, 360, 246–253. http://doi.org/10.1016/j.scitotenv.2005.08.038

Johnes, P. J. (1996). Evaluation and management of the impact of land use change on the nitrogen and phosphorus load delivered to surface waters: The export coefficient modelling approach. Journal of Hydrology, 183(3–4), 323–349. http://doi.org/10.1016/0022-1694(95)02951-6

Johnes, P., Moss, B., & Phillips, G. (1994). Lakes - Classification & Monitoring: A strategy for the classification of lakes.

Lindsay, J. B. (2005). The Terrain Analysis System: A tool for hydro-geomorphic applications. Hydrological Processes, 19(5), 1123–1130. http://doi.org/10.1002/hyp.5818

Line, D. E., White, N. M., Osmond, D. L., Jennings, G. D., Mojonnier, B., Line, D. E., … Mojonnier, C. B. (2015). Pollutant Export from various land uses in the Upper Neuse River basin. Water Environment Federation, 74(1), 100–108. Retrieved from http://www.jstor.org/stable/25045577

Littlewood, I. G. (1992). Estimating contaminant loads in rivers : a review. Walliingford. Oxfordshire, United Kingdom.

Mackereth, F. J. H., Heron, J., & Talling, J. F. (1978). Water Analysis: Some Revised Methods for Limnologists. Scientific Publication No. 36, Ambleside, Cumbria: Freshwater Biological Association.

MacKillop, F. (2012). Climatic city: Two centuries of urban planning and climate science in Manchester (UK) and its region. Cities, 29(4), 244–251. http://doi.org/10.1016/j.cities.2011.10.002

Mainstone, C., Parr, W., & Day, M. (2000). Phosphorus and River Ecology: Tackling Sewage Inputs. Prepared on behalf of English Nature and the Environment Agency.

Manchester City Council. (2007). Report for information. Manchester.

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Marsalek, J., Karamouz, M., Goldenfum, J., & Chocat, B. (2006). Urban water cycle

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processes and interactions (Working Series SC-2006/WS/7). IHP-VI Technical Document in Hydrology UNESCO. Paris, France.

Meybeck, M., & Helmer, R. (1989). The quality of rivers: From pristine stage to global pollution. Palaeogeography, Palaeoclimatology, Palaeoecology, 75(4), 283–309. http://doi.org/10.1016/0031-0182(89)90191-0

Nyamangara, J., Jeke, N., & Rurinda, J. (2013). Long-term nitrate and phosphate loading of river water in the Upper Manyame Catchment, Zimbabwe. Water SA, 39(5), 637–642. http://doi.org/10.4314/wsa.v39i5.7

Rees, A., & White, K. N. (1993). Impact of combined sewer overflows on the water quality of an urban watercourse. Regulated Rivers: Research & Management, 8, 83–94.

Rothwell, J. J., Dise, N. B., Taylor, K. G., Allott, T. E. H., Scholefield, P., Davies, H., & Neal, C. (2010). A spatial and seasonal assessment of river water chemistry across North West England. The Science of the Total Environment, 408(4), 841–55. http://doi.org/10.1016/j.scitotenv.2009.10.041

Walling, D. E., & Webb, B. W. (1985). Estimating the discharge of contaminants to coastal waters by rivers: Some cautionary comments. Marine Pollution Bulletin, 16(1977), 488–492. http://doi.org/10.1016/0025-326X(85)90382-0

Walsh, C. J., Sharpe, A. K., & Breen, P. F. (2001). Effects of Urbanization on Streams of the Melbourne Region, Victoria, Austrailia. I. Benthic Macroinvertebrate Communities. Freshwater Biology, 46, 535–551.

Williams, A. E., Waterfall, R. J., White, K. N., & Hendry, K. (2010). Manchester Ship Canal and Salford Quays: industrial legacy and ecological restoration. In K. Batty, Lesley; Hallberg (Ed.), Ecology of Industrial Pollution (pp. 276–308). Cambridge University Presss.

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Chapter 5 CATEGORISING THE BENTHIC MACROINVERTEBRATE ASSEMBLAGES AND WATER QUALITY IN A HIGHLY URBANISED RIVER

Abstract Previous studies have shown the water quality of urban River Medlock to be good except for PO4-P and a degraded benthic invertebrate community. This study aims to investigate the variables that degrade the benthic macroinvertebrates and so prevent compliance with the EU WFD. Water samples were collected fortnightly over a full season (from March 2013 to April 2014) from five sites and the benthic invertebrate community was sampled monthly. The sample sites were selected upstream and downstream of the WwTW and CSOs over a distance of 17 km. Physico-chemical variables consisting of dissolved oxygen, pH, conductivity, temperature, suspended solids, nutrients, flow and discharge plus the benthic macroinvertebrate community were analysed. The effect of substrate was investigated by the use of colonisation samplers. In this assessment, the use of biotic indices and multivariate statistic tool is presented as an objective tool in the classification of the river. Correlation between the variables using PRIMER-6 BIOENV showed the assemblages of benthic macroinvertebrates were strongly associated with natural variables (altitude, slope and catchment area) plus anthropogenic influenced discharge, conductivity and nutrients. The Medlock was characterised on the basis of benthic invertebrate assemblage into upstream good and downstream poor sites. The poor section had abundant but moderately pollution sensitive taxa dominated by Gammaridae. Seasonal studies revealed a greater abundance of pollution tolerant taxa during winter. This study suggests the Medlock is influenced by the urban stream syndrome and questions the possibility of the Medlock in achieving a good ecological status due to the extent of re-engineering undertaken in the previous two centuries.

Keywords: River Medlock, benthic macroinvertebrates, pollution, urbanisation, PCA, BIOENV, SIMPER

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5.1 Introduction Benthic macroinvertebrate communities and biological indices derived from them have been used in the assessment of running water ecosystems since the

Nineteenth Century (Tate & Heiny 1995; Hellawell 1986; Borja & Franco, 2000). A key reason is that benthic macroinvertebrates are abundant and reasonably sedentary in the absence of a marked change in the physico-chemical environment which allows them to integrate with environmental stress (Paul & Meyer, 2001). Indices such as the

Biological Monitoring Working Party (BMWP) (Hawkes 1997) score and ASPT

(average score per taxon) are commonly used to assess environmental conditions

(Davy-Bowker et al. 2008). A new index based on the BMWP score called the Whalley

Hawkes, Paisley and Trigg (WHPT) metric (Paisley et al. 2014) includes an abundance weighting and inclusion of further taxa, replaced the BMWP score in the

UK in 2014 (Paisley et al. 2014; Environment Agency 2015b). Benthic invertebrates were classified by the UK Environment Agency on the basis of the biological monitoring working party (BMWP) score under the WFD monitoring cycle 1 (2009 -

2015). The new WHPT index is being employed under WFD cycle 2.

The degree at which observed benthic macroinvertebrate community differs from the expected can be predicted using key physico-chemical factors, including hydromorphological variables (Wright et al., 1984). Among the key hydromorphological variables is water discharge. Water discharge plays a key role in invertebrate movement by drift associated with flood conditions and as a result of physical disturbance of the substrate (Brittain & Eikeland 1988).

The Lotic-invertebrate Index for Flow Evaluation (LIFE) score was developed to evaluate benthic invertebrate communities in a river based on the flow regime in a water body (Extence et al. 1999). Discharge and flow will be markedly affected by the increased and highly variable run-off characteristic of urbanised rivers (Lytle & Poff

2004) plus re-engineering of the stream-bed for flood control (Paul & Meyer 2001).

Urbanisation exerts a major effect on water quality due to release of industrial and domestic wastes plus contaminated run-off (Paul & Meyer, 2001). Reduced

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infiltration due to a large impermeable surface area results in a hydraulically efficient drainage system (Walsh et al. 2005) characterised by high runoff velocities, episodic

“flashy” stream flows, increased peak discharge and greater erosion. The changes in the urban river as a result of these conditions are termed, the ‘urban stream syndrome’ (USS) (Walsh et al. 2005). Thus, the term USS has been coined to describe the ecological degradation of urbanised water courses (Walsh et al. 2005). Willemsen et al. (1990) showed that increased episodicity in urban rivers amplifies the change in community structure as the biota has to cope with these variations in velocity

(Nilsson & Renöfält 2008). Therefore, the effects of USS on river quality would impact on compliance with the European Union’s Water Framework Directive (WFD;

Council of the European Union, 2000).

The River Medlock is subject to urban pollution and flooding (Environment

Agency 2009a) and is classified as a “highly modified water body” (Environment

Agency 2009b) subject to pollution from urban runoff, combined sewer overflows

(CSOs) and from wastewater treatment works (WwTWs). As a result the river fails to meet the requirements of the WFD (EA, personal communication 2014). While point source pollution are continuously monitored and regulated, diffuse sources are less controlled and, this has been perceived to have prevented the river from compliance with the WFD (James et al., 2012).

Invertebrate colonisation was investigated on the Medlock in order to determine the effect of substrate on the invertebrate community. Colonisation samples have the advantage of eliminating differences arising from changes in substrate and hence facilitate inter-site comparisons (Davies 2002). The period of exposure of these samplers in freshwater varies from four to six weeks (Weber 1973).

Active sampling which uses the kick-sampling method was applied at all sites while colonisation of invertebrates were employed at two sites upstream and downstream of the WwTWs.

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The aim of this chapter is to determine the relative importance of water quality and quantity (discharge and flow) on the benthic macroinvertebrate community of the River Medlock. A subsidiary aim was to compare the BMWP scores used in the analysis of historic data with the new WHPT index in response to organic pollution. The hypothesis is that physical rather than chemical pollution is the major determinant of ecological status in the river. A second hypothesis is that the new WHPT index would provide better representation of the river’s benthic macroinvertebrates compared to the BMWP scores.

The specific objectives are therefore to:

1. Characterise the benthic macroinvertebrate community in the River Medlock

spatially and temporally;

2. Identify the water quality variables that influence the benthic

macroinvertebrate community;

3. Investigate the relationship between the invertebrate community and selected

physico-chemical factors, including flow rate and discharge;

4. Compare the ASPT and number of taxa from BMWP scores with the WHPT

ASPT and WHPT NTAXA.

5.2 Methodology and Approach

5.2.1 Study area The Mersey catchment is one of the most urbanised catchments in the UK with a catchment size of 4680 km2. The River Medlock (See Figure 1-5) rises in the hills to the NE of Oldham in Greater Manchester (National Grid Reference (NGR): SD 95308

05431), and it flows for 22km, initially through a steep-sided wooded area for 10 km before entering a largely urbanized area and, continues in a SW direction to discharge into the River Irwell immediately downstream of Manchester city centre

(NGR: SJ 85781 97858). The catchment area of the Medlock is 57.5 km2 and the average rate of flow as recorded in the National Rivers Flow Archive was 0.82m³s¯¹

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(CEH, 2016). The majority of the catchment (37%) is heavily urbanised and includes light industries extending into parts of Manchester and the southerly Pennine hills

(CEH, 2016).

Eleven kilometres of the river were examined from Mill Brow Bridge (NGR:

SD 94183 02262) to Pin Mill Brow (NGR: SJ 85781 97858). The rationale is that the catchment of the surveyed reach is largely urbanised. The sub-catchment area of each sample location and other characteristics is shown on Table 5-1. The survey reach of the river has a continuously operational waste water treatment works (WwTW) at

Failsworth (NGR: SJ 89674 99800), about 30 CSOs (EA, personal communication 2016) and an unknown number of surface water drains.

Table 5-1: Catchment area of the sampling sites and WwTw on the Medlock and the distance from the river’s source

Site name Site Catchment Catchment Dist (km) % Altitude Slope No. area (km²) area as % from urban Latitude Longitude (m) (%) of total source cover Mill Brow S1 15 26 6.60 33 53.5173 -2.0892 138.51 2.69 Park Bridge S2 23.5 41 8.50 45 117.87 2.34 53.51282 -2.0997 Road Daisy Nook S3 29.7 52 10.30 42 88.81 2.21 53.50107 -2.12398 Garden Millstream S4 43.9 76 13.00 45 66.81 1.92 53.49258 -2.16317 Lane Purslow S5 53.7 93 16.10 47 47.35 1.67 53.48197 -2.21164 Close Pin Mill S6 54.4 95 17.40 48 42.39 1.57 53.47726 -2.21571 Brow WwTW n/a n/a n/a 12.60 n/a n/a n/a

5.2.2 Sampling and data collection Five sample sites (S1 –S2, S4 to S6) were selected on the river; upstream and downstream of the major WwTW, Failsworth plus major CSOs (United Utilities, personal communication, 2013) and sampled from March 2013 to April 2014 (Table

5-1, Figure 2-2). The presence of riparian vegetation stabilises the river bank and helps to slow down flood water and acts to deposit sediments which serve to build the banks. Failsworth WwTW is situated 12.6 km south of the river’s source and is located between sites 2 and 4. The site numbers S1, S2, S4 to S6 had been retained to concur with the sample sites in previous chapters. Physico-chemical variables were 116

recorded every two weeks but benthic invertebrates were collected monthly since no major change would be expected within two weeks. The water quality data was averaged over the previous two weeks in order to align to the benthic macroinvertebrate samples.

The sub-catchment areas (Table 5-1) were determined using the Terrain

Analysis System, GIS (Lindsay 2005) and a 50m (horizontal) and 0.1m (vertical) digital terrain model (DTM) (Centre for Ecology and Hydrology DTM).

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S1 NGR: SD 94183 02262 S2: NGR: SD 93489

01798

S3 NGR: SD 91874 S4 NGR: SJ 89272 99554

00493

S5 NGR: SJ 86052 98382 S6 NGR: SJ 85781 97858

Figure 5-1: Photographs of sample sites S1 to S6 and the riparian vegetation. S6 shows the debris screen which is aimed to retain large objects and prevent flood damage.

The sample sites on the Medlock as shown on Figure 2-2 and photographs on

Figure 5-1 showed that the river is shallow, the upper sites S1 and S2 are largely erosional while sites S4 to S6 are deeper and depositional with finer substrate. S6 has been canalised for flood control and all sites showed extensive riparian vegetation.

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5.2.2.1 Physical and chemical measurement At each site, measurements of pH, dissolved oxygen, temperature and conductivity were taken using a pre-calibrated hand-held multiparameter water quality meter (YSi 556 Multi probe system YSI, Yellow Springs, Ohio, USA). Water velocity was measured at intervals using the float method i.e. by recording the time taken for the float to travel over a given distance (10m) along the river.

Fortnightly discharge was calculated for each sub-catchment areas on the basis of their relationship to the total catchment area (Table 5-2).

Table 5-2: Average (no.= 23) width (m), depth (m), flow velocity (msˉ¹) and discharge (msˉ¹) at each sample site

Depth (m) Velocity (msˉ¹) Discharge (m³sˉ¹) Site Width Average Min Max Average Min Max Average Min Max S1 5.7 0.22 0.13 0.3 0.27 0.16 0.53 0.15 0.05 0.46 S2 8.2 0.23 0.12 0.59 0.26 0.11 0.59 0.23 0.08 0.72 S4 8.8 0.27 0.14 0.64 0.6 0.14 1.25 0.43 0.15 1.34 S5 8.8 0.29 0.11 0.69 0.57 0.13 1 0.53 0.18 1.63 S6 8.5 0.29 0.15 0.58 0.57 0.13 1 0.53 0.18 1.66

Discharge records obtained at the continuously gauged Environment Agency site was considered a preferable option because discharge could not be accurately calculated due to poor access and high flows. The calculation of discharge data for the sample sites have been described in Chapter 2. Velocity readings were obtained using the float method and confirmed with EA velocity readings obtained from the gauging station described in section of Chapter 2. Table 5-2 showed that average velocity in the river ranged from 0.27ms¯¹ to 0.57ms¯¹ with higher velocity downstream of the WwTW (S4 to S6). Similarly, average discharge from the river showed a range of 0.15m3s-1 to 0.53m3s-1. The results showed that increased volume of water comes from the treatment works flows downstream of the river. At each sampling date a one-litre water sample was collected in an acid-washed polypropylene bottle. A 300ml aliquot was filtered through a 0.45 µm glass fibre paper (VWR International) and oven dried at 105°C for 24 hours to determine total suspended solids. The remainder (700ml) of water sample was filtered through a 119

0.45µm Millipore (Millipore-UK, Limited) hydrophilic, 0.45 µm, 47 mm cellulose acetate filter for measurement of ammonia (ammonia-N), NO3-N (nitrate-N), PO4-P

(phosphate-P) and trace metals. The trace metal sample was acidified with two drops of ultra-pure reagent nitric acid to pH=2 to retain the metals in solution.

A brown glass bottle was used for the collection of samples for measurement of BOD from each site to avoid autotrophic metabolism and incubated at 20°C for five days. The difference between the dissolved oxygen levels was measured using a calibrated Hanna meter (Hanna Instruments Ltd, Bedfordshire, UK) to determine the

BOD5 recorded on the first day and the fifth day was calculated to determine the

-1 BOD5 (mgL ).

The trace metals chromium (Cr), cadmium (Cd), copper (Cu), nickel (Ni), lead

(Pb) and zinc (Zn) were analysed by Inductively Coupled Plasma-Mass Spectrometry

(ICP-MS) using an Agilent 7500cx (Agilent Technologies, Santa Clara, USA) spectrometer. Calibration was by matrix-matched standards.

5.2.2.2 River Substrate River substrate was characterised by estimating the percentage substrate contribution (Table 5-3) using the Wentworth scale. The highest contribution was from sand which accounted for 36% of the total. The sites downstream have higher finer substrates.

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Table 5-3: Types of substrate found at sample sites. *The text indicated in bold showed the dominant substrates to be stones and sand.

Site Boulders Stones Pebbles Gravel Sand Silt Mud (%) (%) (%) (%) (%) (%) (%)

S1 6 37 7 5 30 15 3.5 S2 10 39 7 7 27 12 3 S4 4 27 1 23 32 12 3 S5 14 43 3 6 28 8 0 S6 6 18 1.4 3.8 42 16 2 % total 7.5 30.84 3.65 8.43 35.55 11.85 2.16 contribution

5.2.2.3 Nutrient analysis

NO3 and PO4 samples were processed within 24 hours of sample collection and analysed using a SEAL Auto Analyzer 3 High Resolution instrument (SEAL

Analytical Ltd, Southampton) based on a segmented flow analysis. For further information on the methods employed by the autoanalyser see SEAL Analytical

(2013). Throughout this study, concentrations of nutrients are presented as elemental concentrations, i.e. mgLˉ¹ P, not PO4 and mgLˉ¹ N, not of NO3; ammonia as mgLˉ¹ of

N not NH3.

Analysis of PO4-P is based on the molybdenum blue method in which orthophosphate reacts with molybdate and ascorbic acid to form an intensely blue compound which is measured at 660nm. Detection limits for PO4-P measured as P was 0.004 mgLˉ¹.

Nitrate (NO3-N), measured as N following the DIN 38405 and ISO/DIS 13359 standard methods and with a detection limit of 0.01 mgLˉ¹. The analysis was based on the cadmium reaction method in which the sample is reduced from nitrate-N to nitrite by hydrazine in alkaline solution, with a copper catalyst, after which it is, reacted with sulfanilamide and coupled with α-napthyethlene diamine dihydrochloride to form a pink compound measured at 520 nm. Phosphoric acid is added at the final stage to reduce the pH and thus avoiding precipitation of calcium

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and magnesium hydroxide. The addition of zinc to the reducing agent suppresses the complexing of copper by organic material. These analyses comply with the Standard

Committee of Analysts Publications, (2011).

Ammonia-N concentration (mgL¯¹) was analysed using the Hanna (Hanna

Instruments Ltd, Leighton Buzzard, Bedfordshire, low range reagents (HI-93700-01) kit by spectrophotometry. The limit of detection for ammonia-N measured as N was

0.01mgLˉ¹. The analysis of ammonia-N is based on the Nessler Method in which the

Nessler Reagent (K2HgI4) reacts with the ammonia-N present in the sample under strongly alkaline conditions to produce a yellow-coloured species. The intensity of the colour is in direct proportion to concentration of ammonia-N concentration. The measurement wavelength is 425 nm.

Geographical information including sub-catchment altitude was obtained from the internet map tools (www.freemaptools, www.daftlogic.com) for the study sites using the sub-catchment latitude and longitude. The sub-catchment slopes were obtained by dividing each site’s elevation from the river’s source by the distance of the sample site from the source.

5.2.2.4 Benthic invertebrates sampling and collection Samples were collected from five sites (S1, S2, S4 to S6) using a 1mm mesh hand net by the three-minute kick net sampling technique outlined in the Water

Framework Directive, UK policy report (UK Technical Advisory Group 2008). An additional one-minute manual search was carried out by collecting benthic invertebrates that could have been missed through the kick sampling. The samples were preserved in 70% ethanol, identified and counted in the laboratory (Pawley,

Dobson, & Fletcher, 2014). Biotic indices used in this study were the Biological

Monitoring Working Party (BMWP), Average Score Per Taxa (ASPT), Whalley,

Hawkes, Paisley & Trigg (WHPT), River Invertebrate Classification Tool (RICT) for

Environmental Quality Ratio and, Lotic Invertebrate-Index Flow Evaluation (LIFE).

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The BMWP score and WHPT were used simultaneously in this study because both indices fall under WFD reporting period encompassed by this study. While the

BMWP scores is still used for local EA operations as the scores can easily be communicated to and understood by members of the public, this is not yet the case with WHPT index. The WHPT index is currently rarely used for river classification due to the quantification required and therefore, only applied in EA reports meant for reporting for specialised audience. The EQR from the RICT is derived from both indices and provides an overall WFD status for a site. This study falls within the two reporting cycles of the WFD and therefore applied in this study to compare and contrast output of the ASPT and NTAXA.

Under the BMWP score, all macroinvertebrates apart from Oligochaeta were identified to family level using the taxonomic groups used in the biological monitoring working party (BMWP)-score (Hawkes 1997). Invertebrate families that are very sensitive to sewage pollution receive scores of 10 and the most tolerant families receive a score of 1 and the sum of the total scores determine the category to which the river is classified as shown on Table 5-4.

Table 5-4: BMWP Scores and interpretation (Hawkes 1997).

The Whalley Hawkes, Paisley and Trigg (WHPT) Average Score Per taxa

(WHPT ASPT) and WHPT number of taxa (WHPT NTaxa) were assessed separately and then combined in a “worst of” approach to provide the overall invertebrate classification. The WHPT ASPT was applied as abundance weighted metric. Table

5-5 shows WHPT logarithmic abundance categories and Environmental Quality

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Ratio (EQR). Thus, the sum of the total scores from the samples collected will determine the category to which the river is classified.

Table 5-5: WHPT logarithmic abundance categories and Environmental Quality Ratio (EQR) for WHPT-ASPT and WHPT-NTAXA.

Abundance Numerical category Abundance WHPT EQR NTAXA ASPT AB1 1-9 High/Good 0.80 0.97 AB2 10-99 Good/Moderate 0.68 0.87 AB3 100-999 Moderate/Poor 0.56 0.72 AB4 >1000 Poor/Bad 0.47 0.59

The UK River Invertebrate Classification Tool (RICT) is used to contextualise

WHPT scores by predicting site specific reference values and provides a WFD compliant probabilistic classification (Davy-Bowker et al. 2008). The ecological quality of the river is classified for spring and autumn seasons.

The ASPT of the samples observed (Obs) is divided by the predicted (Pred) pristine condition score to provide a classification Ecological Quality Ratio (EQR) belonging to any of the WFD classes as shown on Table 5-5. EQR values close to 1 indicate invertebrate communities close to the natural state, those near to zero indicate a high level of pollution or disturbance.

LIFE scores are determined by using the family abundance and ecological associations with flow as shown on Table 5-6. The sum of the individual invertebrate families is divided by the number of scoring families to produce the overall LIFE Score: LIFE =

LIFE scores less than 6.00 generally indicate sluggish or still water conditions.

As current velocity increases, so do LIFE scores. LIFE values greater than 7.5 indicate very fast flows.

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Table 5-6: Life Flow Groups.

ABUNDANCE CATEGORY LIFE Flow Velocity A B C D/E Flow association (cmsˉ¹) (1-9) (10-99) (100-999) (1000 - > Group 10,000) I Rapid flows Typically 9 10 11 12 >100cmsˉ¹ II Moderate to 20 cmsˉ¹ to 8 9 10 11 fast flows >100cmsˉ¹ III Slow to < 20cmsˉ¹ 7 7 7 7 sluggish IV Slow and N/A 6 5 4 3 standing V Standing N/A 5 4 3 2 water VI Drying and N/A 4 3 2 1 drought impacted

5.2.2.5 Colonisation samplers The colonisation method was used in addition to the kick sampling (active method) in order to allow comparison between sites as colonisation is independent of the natural substrate (Czerniawska-Kusza 2004). Colonisation will therefore facilitate examination of the impact of stream substrate on the community. Two artificial colonisation samplers (Figure 5-2) were positioned at two locations, upper S2 and lower S6 for a thirty-day period over four month duration from September 2014 to

December 2014. The 30-day period was critical for the development of a representative community of organisms (Weber 1973; Meier et al. 1979). The two sites were selected in order to determine if the substrates impacted on the benthic invertebrate community as S2 is largely erosive and S6 is partly depositional, having a sandier substrate. Both sites are located upstream and downstream of the major

WwTW respectively. At the end of the 30 day period the colonisers were removed from the river and washed in a bucket to be processed at the laboratory using the same procedure as the kick samplings.

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Figure 5-2: Invertebrate colonisation sampler before and after 30 days’ colonisation. Source, Author, 2014

5.2.2.6 Statistical analysis Data were tabulated and analysed using Microsoft Excel 2013, GraphPad

Prism 6 and multivariate analysis (similarity percentages (SIMPER) routine; Principal component analysis (PCA); Non-metric multidimensional scaling (nMDS); BIOENV

(Biota and Environmental) procedure were performed using the PRIMER-6 software package (Clarke & Warwick 2001). Multivariate approaches (Cao et al. 1996) were used to identify the water quality variables that affect the macroinvertebrate community in the River Medlock. The study variables include in addition to benthic macroinvertebrate assemblages, physico-chemical parameters collected at a number of sites along the river over a full season between March 2013 and April 2014.

Differences between the physico-chemical variables and the biotic indices

(BMWP, ASPT, and WHPT) at each site were analysed using One-way Analysis of

Variance (ANOVA). Pearson correlation analysis was used to investigate how the various metrics changed and how these variables impacted on the river.

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5.3 Results The following results are presented in three sections- physico-chemical variables, benthic macroinvertebrates and the relationship between physico-chemical parameters and benthic macroinvertebrates. The rational is because conventionally, physico-chemical variables are presented first and the water quality would need to be established in order to determine if the biota will correspond to the physico-chemical parameters. The description of each parameter in the results is referred in terms of sample location and distance from source.

5.3.1 Physical and chemical variables The spatial differences and ordination of the physico-chemical variables were determined. On the basis of the Gregorian calendar used in separating the seasons-

December, January and February (Winter); March, April and May (Spring); June, July and August (Summer); September, October and November (Autumn), there was no significant (p>0.05) difference between the water quality variables (including DO, pH, temperature, conductivity, BOD, Ammonia, NO3-N and PO4-P) with season and is therefore not presented.

5.3.3.1 Ordination of environmental variables Ordination of variables was described in two scenarios in order to determine the differences in outcomes between using all or fewer environmental variables in the

PCA. All data matrix were transformed and ordinations were executed on the basis of a distance matrix.

Scenario 1: Twenty physico-chemical variables, specifically dissolved oxygen, pH, temperature, conductivity, suspended solids, BOD, ammonia-N, NO3-N, PO4-P, discharge, velocity, catchment area, altitude, slope, boulders, stones, pebbles, gravel, sand and silt were combined in a PCA. (Figure 5-3, Table 5-7). The first PCA axis accounted for 34% of the overall variance and was most heavily weighted to altitude, slope, catchment area, PO4-P, NO3-N and velocity while the second PC axis accounted for 17% of the variance and was dominated by substrate, specifically 127

boulders, stones, sand and silt. The PCA plot representing PC1 and PC2 showed a difference between the sites especially with reference to the river’s physical attributes.

Table 5-7: All environmental variables (p<0.05) based on ordination with principal components. Selected PC characters in bold indicated environmental variables that controlled the river.

Variable PC1 PC2 PC3 PC4 PC5 DO -0.072 -0.06 0.231 -0.009 -0.562 PH 0.141 -0.081 -0.295 -0.373 -0.282 TEMP 0.022 -0.001 -0.444 0.39 0.287 COND 0.115 0.02 0.333 -0.269 0.221 BOD 0.138 0.025 -0.024 -0.234 0.033

NH3-N 0.045 0.081 0.42 0.222 0.364

NO3-N 0.304 0.008 -0.259 0.068 0.133

PO4-P 0.318 -0.007 -0.208 0.154 0.001 SS 0.148 0.18 0.185 0.314 0.169 DISCHARGE 0.212 0.116 0.451 -0.004 -0.075 VELOCITY 0.239 0.068 -0.007 0.174 -0.263 CATCHMENT AREA 0.372 0.071 -0.001 -0.094 -0.008 BOULDERS -0.007 0.485 -0.074 -0.238 0.116 STONES -0.168 0.474 -0.071 0.032 -0.1 PEBBLES -0.356 0.124 -0.055 -0.117 0.107 GRAVEL 0.136 -0.098 0.077 0.475 -0.35 SAND 0.15 -0.445 0.05 -0.169 0.199 SILT -0.144 -0.487 0.059 -0.009 0.137 ALTITUDE -0.368 -0.051 0.008 0.132 -0.014 SLOPE -0.364 -0.061 0.002 0.151 -0.047

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4 Site Number 1 2 4 5 6

2 STONESBOULDERS

SS (17%) PEBBLES DISCHARGE NH3-NVELOCITYCATCHMENT AREA CONDBOD 2 NO3-N

C 0

P TEMP PO4-P ALTITUDESLOPE DO GRAVELPH

SAND SILT -2

-4 -6 -4 -2 0 2 4 PC1 (34%)

Figure 5-3: Ordination diagram of 20 environmental variables at each of the sites on the River Medlock. Variables are DO (Dissolved oxygen), pH, temperature (TEMP), conductivity (COND),

velocity, BOD, ammonia-N, NO3-N, PO4-P, SS (suspended solids), discharge, substrates, catchment area, altitude and slope.

Scenario 2: In order to determine impact of water quality variables, altitude, slope,

catchment area and substrates were excluded in a further PCA combination. The first

PCA axis accounted for 26% of the overall variance and was most heavily weighted

on NO3-N, PO4-P, velocity and suspended solids. The second axis accounted for 21%

of the variance and was dominated by discharge, ammonia-N and conductivity

(COND) (Figure 5-4, Table 5-8). These groups indicated the importance of water

quality on river assessment.

The output for scenarios 1 and 2 were similar especially in the distinction

between the sample sites S1 and S2; and, S4 to S6. The exclusion of certain variables

in the second scenario showed that the nutrients and suspended solids concentration

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was influenced by velocity and discharge which would be increased with decreasing slope.

4 Site Number 1 2 4 5 6 2 DISCHARGE NH3-N (21%) COND SS

VELOCITY

2 BOD

C 0

PO4-P NO3-N

TEMP

-2

-4 -4 -2 0 2 4 PC1 (26%) Figure 5-4: Ordination diagram of 11 environmental variables at each of the sites on the River Medlock. Variables are DO (Dissolved oxygen), pH, temperature (TEMP), conductivity (COND), velocity, BOD, ammonia-N, NO3-N, PO4-P, SS (suspended solids) and discharge.

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Table 5-8: Selected environmental variables (p<0.05) based on ordination with principal components. PC characters in bold indicated the water quality variables including nutrients and DO influenced the river.

Variable PC1 PC2 PC3 PC4 PC5 DO -0.208 0.163 0.165 0.531 0.095 pH 0.217 -0.284 0.549 0.076 -0.119 TEMP 0.142 -0.411 -0.477 -0.171 0.075 COND 0.148 0.365 0.300 -0.316 -0.466 BOD 0.246 0.02 0.411 -0.384 0.543 Ammonia-N 0.071 0.455 -0.267 -0.325 0.415

NO3-N 0.521 -0.172 -0.012 -0.031 -0.095

PO4-P 0.516 -0.132 -0.045 0.117 -0.017 SS 0.280 0.273 -0.316 -0.003 -0.433 DISCHARGE 0.246 0.508 0.041 0.233 0.069 VELOCITY 0.354 0.055 -0.100 0.512 0.292

5.3.3.2 Spatial differences at sample locations Significant (p<0.05) differences between S1, S2 and from S4 to S6 were observed for pH, conductivity, discharge, NO3-N and PO4-P (Figure 5-5,Table 5-10).

Low conductivity upstream indicated that the river’s conductivity is not influenced by the river’s geology (and from the moors) which is made up mainly of mixed permeability superficial deposits. As expected, discharge increased significantly downstream of the river along with increased concentration of nutrients from sites S4 to S6 (Figure 5-5; D&E). The change occurred from S4 (which is 0.5km below the

WwTW) to S6 as all three sites are located downstream of the WwTW. No significant

(p>0.05) difference was recorded for temperature, BOD, ammonia-N, dissolved oxygen and suspended solids. However, periods of high concentrations of suspended solids were observed from S4 to S6 and this was linked to increased discharge from runoff and CSOs especially during high rainfall events. Post Hoc, LSD for substrates indicates that sand and stones were abundant in the river compared to other substrates (p < 0.001, F6, 28= 24.963).

Within the context of the Water Framework Directive, all average chemical variables fulfilled the WFD requirements except PO4-P (European Union, 2000). Mean suspended solids concentration conformed to the requirement of the Freshwater

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Fisheries Directive even though there were periods (probably during storm

conditions) when the concentration of suspended solids exceeded the standard of

25mgLˉ¹.

Trace metals were below environmental quality standards as shown on Table

5-9. They are therefore highly unlikely to influence benthic invertebrate community

structure.

Table 5-9: Mean and standard deviation of trace metals sampled at six sites on the River Medlock between April 2013 and April 2014.

Trace metal Cr Cu Cr Cr Cr standard standard standard Cr standard standard standard standard (ug/L) (50 ug/L) (50 ug/L) (3000 ug/L) (5 ug/L) (1 ug/L) (50 ug/L)

Number of Cu Cd Hg Sites samples Cr (ug/L) (ug/L) Zn (ug/L) (ug/L) (ug/L) Pb (ug/L) Ni (ug/L) S1 12 0.24±0.12 6.59±4.01 24.26±12.48 0.03±0.01 0.07±0.0 0.91±1.53 2.45±0.69 S2 12 0.25±0.12 7.09±3.22 24.95±17.22 0.03±0.01 0.07±0.01 0.52±0.39 2.85±0.94 S3 12 0.30±0.17 7.49±4.42 34.43±45.41 0.03±0.01 0.08±0.01 0.55±0.27 2.62±0.52 S4 12 0.39±0.21 9.55±4.26 44.15±62.47 0.03±0.01 0.07±0.00 0.68±0.89 2.71±0.38 S5 12 0.33±0.13 9.40±3.32 22.92±13.89 0.03±0.01 0.07±0.00 0.38±0.20 2.56±0.40 S6 12 0.35±0.10 8.94±2.39 21.48±10.79 0.04±0.02 0.07±0.00 0.61±0.51 2.62±0.40

The influence of land use on the five sampling sites was examined by

comparing the percentage size of the sub-catchment areas. The results showed S1 had

significantly (p <0.05) low percentage land use compared to other sites along the

river.

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Figure 5-5: (A) Discharge, (B) Suspended solids, (C) Conductivity, (D) pH, (E) NO3-N and (F) PO4-P at all sample locations on the River Medlock. Box and whiskers represent 25% and 75%, median, minimum and maximum values of the variable measured. The dotted lines show the WFD standards.

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Table 5-10: One-way ANOVA to compare sites S1 and S2, S4 to S6 for mean environmental variables, BMWP, WHPT scores and ASPT of benthic invertebrates and trace metals. The results are compared with European Union’s standard requirements.

Variables ANOVA P ( post hoc Average Standards Interpretation test) conc. (WFD, FFD)

Discharge F4, 110 = 7.061 *** S1-S2 & 0.37 (m³s ˉ¹) S4-S6 Conductivity F4, 103 = 8.081 *** S1& S3- 615.30 N/A (µScmˉ¹) S6; ***S2 & S3-S6 pH F4, 108 = 3.232 ***S2 & S4- 8.03 S6 Good Temperature F4, 108 = 0.324 ns 10.13 (°C) DO (%sat) F4, 108 = 0.899 ns 101.58 >80% High Suspended F4,108 = 1.633 ns 9.81 ≤25mgLˉ¹ Good (FFD) solids (mgL ˉ¹) NO3-N F4, 110 =14.34 **S1-S2 & S4 2.89 (mgL ˉ¹) –S6 Ammonia-N F4, 108 = 0.379 ns 0.52 <0.6 mgL¯¹ Good (mgL ˉ¹) PO4-P F4, 110 = 13.32 *** S1&S4 & 0.33 <1mgL¯¹ Poor (mgL ˉ¹) ***S1 &S6 BOD5 F4, 108 = 0.879 ns 2.54 <5 mgL¯¹ High (mgL ˉ¹) Trace metals ns Below Very Good (Cr, Cu, Zn, Cd, environmental Ni, Pb) (µgL ˉ¹) quality standards/detec tion limits Where p < .0001***; p < .05*; not significant (ns)

Correlation analysis showed a significant positive (p < .05) relationship between most of the variables including NO3-N and PO4-P. Discharge correlated with most variables suggesting its key influence on this system.

The maximum and minimum velocity (in cms¯¹) (Table 5-11) recorded in the river fortnightly from March 2013 to April 2014 were estimated from instantaneous velocity readings at the gauging station during the study period and related to each site’s-catchment velocity. This aimed to establish the maximum and minimum velocity that would influence the movement of substrate in the river. By using the

Hjulström-Sundborg diagram (Earle 2015), the particle size distribution was 134

estimated on the basis of the velocity conditions, under which sediment was either eroded, transported or deposited. The results showed that at maximum velocity

(52.63 -100 cmsˉ¹), particles ≤ 0.1mm e.g. silt are transported in suspension while at minimum velocity (11.36 -15.72cmsˉ¹; particles ≤ 1 mm e.g. sand are transported and particles ≤ 2mm e.g. gravel are deposited as bed load. These results suggest an unstable sediment regime which is likely to affect the community of benthic macroinvertebrates.

Table 5-11: Maximum and minimum velocity recorded at the Medlock with for fortnightly data and at the sample locations and instantaneous readings from the EA gauging station (*EA data indicates that records were not obtained during the period).

Sites Maximum Velocity Minimum Velocity (cmsˉ¹) (cmsˉ¹)

S1 52.63 15.72 S2 58.82 11.36 S4 100.00 0* S5 100.00 13.33 S6 66.67 0*

5.3.2 Benthic macroinvertebrates

5.3.2.1 Community structure A total of 32 benthic macroinvertebrate families were recorded in the River

Medlock at the five sites between March 2013 and April 2014 where benthic invertebrates could be safely collected. The benthic macroinvertebrate families were distributed across three phyla (Annelida, Arthropoda and Mollusca). The phylum

Arthropoda was important as they contributed 23 benthic invertebrate families which made up 72% of the total identified at the river. The most common taxa which occurred at all the sites were largely insects (Baetidae, 19%; Chironomidae, 17%) and

Annelids (Lumbriculidae, 14%, Tubificidae, 38%).

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5.3.2.2 Spatial analysis of assemblage structure Benthic macroinvertebrate assemblages in the river were analysed using the

Non-metric multidimensional scaling (nMDS) and Similarity Percentages (SIMPER) multivariate tests. nMDS plot in Figure 5-6 indicated that sites S1 and S2 were similar in composition and, sites S4 to S6 also grouped together.

2D Stress: 0.27 Site Number 1 2 4 5 6

Figure 5-6: MDS ordination plots of benthic invertebrates at sample sites 1-2, 4-6 grouped together.

Similarity percentages (SIMPER) were used to distinguish invertebrate families that made the greatest contribution to the differences identified by the ordination plots in Table 5-12. The results from SIMPER analysis of the invertebrate assemblages showed that both S1 and S2 were similar at 44% while between S4, S5 and S6 the percentage similarity was 53%. The highest average dissimilarity between all sample sites was observed for S1 and S5 at 64.71% (Table 5-12). This difference could be associated with the distance between the upper S1 and lower S5 and the difference in urban extent and differences in altitude. While S1 had an urban area of

33%, 6.6km from source, S5 had an urban area of 47% and 16.1km from source. Also

S1 had more pollution sensitive taxa including Heptageniidae, Leuctridae,

Ephemerellidae and Perlodidae compared to S5.

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Table 5-12: The top-ranked SIMPER contributors to % dissimilarity in benthic macro invertebrate composition between S1 and S5. Figures in bold were highlighted based on the high ranking pollution sensitive taxonomic groups.

Average dissimilarity = 64.71% Site and distance Site 1 Site 5 from source (6.60km) (16.10km) Average Average Cumulative No. Benthic Abundance Abundance Contribution contribution invertebrates (%) (%) 1 Gammaridae 0.19 1.65 12.19 12.19 2 Heptageniidae 1.38 0.47 8.78 20.97 3 Chironomidae 1.15 0.98 8.02 28.98 4 Simuliidae 0.81 0.19 5.99 34.97 5 Tubificidae 0.84 0.07 5.94 40.92 6 Paediciidae 0.79 0.11 5.66 46.58 7 Lumbriculidae 0.69 0.45 4.77 51.34 8 Leuctridae 0.41 0 4.69 56.03 9 Baetidae 1.41 1.73 4.65 60.69 10 Ephemerellidae 0.54 0.07 4.64 65.32 11 Hydropsychidae 0.35 0.42 4.29 69.61 12 Perlodidae 0.46 0 3.99 73.6 13 Lumbricidae 0.38 0.47 3.85 77.45 14 Tipulidae 0.61 0.47 3.6 81.06 15 Erpobdellidae 0.25 0.56 3.23 84.29 16 Asselidae 0.27 0.30 2.99 87.28 17 Rhyacophilidae 0.19 0.24 2.33 89.61 18 Limnephilidae 0.08 0.21 2.3 91.91

5.3.2.3 Temporal variation in benthic invertebrates and biotic scores Temporal variation in the benthic macroinvertebrates was analysed using the multidimensional scaling ordination (MDS) plot as shown on Figure 5-7. The MDS plot showed slight differences between the seasons. The highest average dissimilarity in the assemblage of benthic invertebrate families occurred between summer and winter at 61.62%, based on similarity percentage (SIMPER) analysis (Table 5-13) while autumn and spring were more alike suggesting a low degree of intra-seasonal variability in invertebrate community. Tubificidae, Tipulidae, Lumbriculidae,

Heptageniidae, Simuliidae, Hydropsychidae and Paediciidae were more abundant in

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the winter than in the summer. Baetidae and Gammaridae were abundant in the summer.

2D Stress: 0.26 Season Sp Su A W

Figure 5-7: MDS ordination plot of benthic invertebrates based on seasonal patterns of spring (Sp), summer (Su), autumn (Au) and winter (W).

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Table 5-13: Assemblage of benthic macroinvertebrates between the summer and winter detected by SIMPER. Figures in bold showed the taxa which dominated the river based on the average abundance.

Average dissimilarity = 61.62% Summer Winter (June – (December – August) February) No Benthic Average Average % contribution Cum.% invertebrates Abundance Abundance 1 Tubificidae 0.20 1.02 8.95 8.95 2 Tipulidae 0.18 0.89 7.97 16.92 3 Lumbriculidae 0.21 0.83 7.41 24.33 4 Heptageniidae 0.38 1.14 7.31 31.65 5 Baetidae 1.67 1.18 7.16 38.81 6 Chironomidae 1.11 1.57 7.10 45.90 7 Simuliidae 0.18 0.81 5.64 51.54 8 Hydropsychidae 0.34 0.60 5.53 57.07 9 Paediciidae 0.28 0.66 4.67 61.74 10 Erpobdellidae 0.12 0.49 4.61 66.35 11 Ephemerellidae 0.67 0.07 4.60 70.95 12 Gammaridae 0.80 0.58 3.93 74.88 13 Asselidae 0.17 0.38 3.57 78.46 14 Lumbricidae 0.15 0.28 3.52 81.97 15 Rhyacophilidae 0.18 0.34 3.43 85.40 16 Leuctridae 0.45 0.00 3.29 88.70 17 Limnephilidae 0.30 0.07 2.70 91.40

Classification of the river using the River Invertebrate Classification Tool

(RICT) during spring and autumn showed that there was a decline in the number of taxa at the study sites (Table 5-14) even though no significant (p>0.05) difference was found between both seasons. Based on the number of taxa in the river, the

Environmental Quality Ratio (EQR) classified S1 as “Good” with a value of 0.73,

“Moderate” at S2 (with a value of 0.63) and “Bad” from S4 to S6 (with a value <0.42).

In order for the river to achieve a high/good category, the number of taxa (NTAXA) must be close to an EQR of 0.8. Classification which was based on the average score per taxon (ASPT) weighed abundance metric indicated the river to be “moderately polluted” at all the sites except S5 which was “Poor”.

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Table 5-14: Biological classification results with environmental quality ratio (EQR)

Indices EQR Status Sites Indices EQR Status ASPT 0.85 Moderate S1 NTAXA 0.73 Good ASPT 0.86 Moderate S2 NTAXA 0.63 Moderate ASPT 0.78 Moderate S4 NTAXA 0.41 Bad ASPT 0.75 Poor S5 NTAXA 0.35 Bad ASPT 0.76 Moderate S6 NTAXA 0.35 Bad

5.3.2.4 BMWP, WHPT, ASPT and LIFE scores BMWP scores and ASPT declined downstream of the river (Figure 5-8). A one- way ANOVA showed a significant (p < .05) difference between the sites S1 and S6 for the BMWP score (F4, 63= 3.889) and for ASPT (F4, 63 = 3.513). The average BMWP scores and ASPT did not exceed 40 and 4.5 indicating the river to be polluted or otherwise impacted (Hawkes 1998).

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S a m p le L o c a tio n s S a m p le L o c a tio n s

Figure 5-8: Box and whisker plot with 25% and 75%, median, minimum and maximum values of BMWP score (A) and ASPT (B) with distance along River Medlock

While a significant (p<0.05) difference was found between the sample sites for WHPT

ASPT (F4, 62 =5.60), there was no difference for the number of taxa (Figure 5-9).

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1 0 1 4 B

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Figure 5-9: Box and whisker plot with 25% and 75%, median, minimum and maximum values for monthly samples obtained between March 2013 and April 2014 of (A) WHPT ASPT (B), (B) WHPT NTaxa with distance along the river Medlock

Comparative analysis of BMWP ASPT, number of taxa with WHPT ASPT and

WHTP NTAXA (Figure 5-10) showed a significant (p<0.05) correlation. ASPT

(r=0.877, Pearson correlation) and the number of taxa showed an almost perfect relationship. This result indicates that either BMWP or WHPT ASPT or NTAXA could equally be used in the classification of invertebrates. The comparison of the two indices for tributaries of River Swale’s metal impacted catchment showed some similarities (Barber 2014).

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W H P T N T a x a W H P T A S P T

Figure 5-10: Comparison of WHPT and BMWP number of taxa and ASPT

The results of LIFE index (Table 5-15) suggested the river to be fast flowing, having an overall LIFE score average of 7.5. This result might be expected given the maximum and minimum velocities recorded during the survey and suggests moderate to high flow conditions. 141

Table 5-15: LIFE results summary

Sample sites n taxa LIFE

S1 147 18 8.17 S2 146 23 6.35 S4 137 17 8.06 S5 129 16 8.06 S6 90 11 8.18

5.3.2.5 Composition of the benthic macroinvertebrate community in colonisation samplers A total of 12 macroinvertebrate taxa were found in the colonisation samplers deployed at S2 (4km above the WwTW) and S6 (5km below the WwTW). After each

30-day period over a four-month duration, the colonisation samplers were dominated by the crustacean family Gammaridae (67%), followed by the insects

Chironomidae (14%) and Hydropsychidae (8%) plus a further crustacean, Asellidae

(5%) (Figure 5-11).

Gammaridae was more abundant at the lower site S6 compared to upstream

S2 which was expected to be in the cleaner part of the river. Kick sampling results also showed that Gammaridae was abundant at the downstream sites. The results suggest that a difference in substrate was not the reason for the abundance of

Gammaridae.

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Figure 5-11: Benthic invertebrate community composition in the invertebrate colonisation samplers at S2 (4km upstream of the WwTW) and S6 (5km downstream of the WwTW).

5.3.3 Relationship between physico-chemical, hydrogeomorphological variables and benthic macroinvertebrate assemblages All environmental variables combined in a PC matrix (Table 5-7) were combined with benthic macroinvertebrate assemblages in the BIOENV analysis to determine which variable(s) affected benthic invertebrate abundance and distribution. BIOENV analysis revealed that the most important variables structuring benthic macroinvertebrate communities were conductivity, PO4-P, discharge, catchment area, altitude and slope based on correlation matrix ρ = 0.274 (Table 5-16).

Several correlation analyses were performed with fewer variables to test if there were

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any changes in the BIOENV outcome. With selected variables from Table 5-8, the most important variables were conductivity, NO3-N, PO4-P and discharge, (ρ = 0.225).

Table 5-16: Correlation between physico-chemical variables and benthic invertebrate assemblages using the BIOENV procedure. The correlation was carried out using a series of different number of variables.

Weighted No. of Correlations variables ρ Selections of variables Conductivity, Discharge, Catchment area, 0.274 5 Altitude, Slope Conductivity, Discharge, Catchment area, 0.273 4 Altitude

Conductivity, PO4-P,Discharge, Catchment 0.272 5 area, Altitude 0.272 2 Conductivity, Catchment area Conductivity, NO3-N, Discharge, Catchment 0.271 5 area, Altitude Temperature, Conductivity, Discharge, 0.271 5 Catchment area, Altitude Conductivity, PO4-P, Discharge, Catchment 0.269 5 area, Slope DO, Conductivity, Discharge, Catchment area, 0.268 5 Altitude Conductivity, NO3-N, Discharge, Catchment 0.268 5 area, slope Conductivity, Discharge, Catchment area, 0.267 4 Slope

5.4 Summary of results

• Dissolved oxygen, pH, BOD, ammonia-N, NO3-N showed that the water

quality is “Good”/ “Low pollution” except PO4-P which was >0.1mg/L and

classifies the river as “poor”. However the benthic invertebrate classification

indicates the river to be “moderately polluted”.

• Spatial variation for physico-chemical and benthic macroinvertebrates showed

the upstream sites (S1 and S2) of the WwTW were better than downstream

sites (S4 to S6). Gammaridae, a taxa found in moderately polluted rivers, was

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found to be dominant at downstream sites especially S5 for (kick sampling)

and S6 (colonisation samplers) even though BMWP scores indicated pollution.

• There was a strong seasonal dissimilarity in benthic macroinvertebrate

abundance between winter and summer. While Oligochaeta, Chironomidae

and Simulidae were abundant during winter, Gammaridae was abundant

during summer.

• The relationship between physico-chemical variables and benthic

invertebrates using the BIOENV analysis showed conductivity, discharge,

catchment area, altitude, slope and nutrients as factors which described the

differences between the study sites. While altitude, slope and catchment area

could be related to the location of the sampling sites, the concentration of

conductivity and nutrients in the river is related to increase in discharge.

• The LIFE index indicates that the benthic macroinvertebrate assemblages were

affected by variation in flow and therefore probably affect the sediment

instability.

• The NTAXA and ASPT from BMWP were compared with WHPT NTAXA and

ASPT and the results revealed no (p>0.05) difference between the indices.

5.5. Discussion The aim of this chapter was to determine the relative importance of water quality and quantity (discharge and flow) on the benthic macroinvertebrate community of the River Medlock. The results of macroinvertebrate sampling from

March 2013 to April 2014 and the colonisation sampling which took place from

September to December 2014 revealed that the Medlock was a moderately polluted system on the basis of fewer numbers of pollution sensitive taxonomic groups.

Various studies of benthic invertebrates in urban rivers have revealed similar patterns of fewer or absent pollution sensitive taxa including Beavan et al. (2001) in the River Tame catchment, UK and worldwide from studies in Brazil, USA, Australia and Canada (Grapentine et al. 2004; Guimaràes et al. 2009; Mikalsen 1989; Silveira et al. 2006; Walsh et al. 2001; Wright et al. 2007; Whitehurst & Lindsey 1990). These 145

studies have all associated faunal impoverishment, degradation of the river and the loss of sensitive taxa to urbanisation rather than water quality. Thus, the loss of sensitive invertebrates, biological degradation in other urban catchments (Duda et al.

1982) plus the Medlock are part of the complex changes in urban rivers resulting from a flashy hydrograph and altered channel morphology collectively termed the

“urban stream syndrome”(USS) (Walsh et al. 2005; Walsh et al. 2012). The most common taxa recorded at all sites throughout the study were Baetidae,

Chironomidae, Lumbriculidae, and Tubificidae. While these taxonomic groups tolerate organic pollution, some researchers have also attributed their dominance to the deposition of silt and sand substrate arising from episodic discharge (Macan 1962;

Chutter 1969; Langford & Bray 1969).

There are a number of possible reasons for the less degraded invertebrate community at the upper sites. S1 has 33% average sub-catchment urban area, whereas the average at S4 to S6 was 45% and are classified as “heavily modified waterbody” (HMWB) by the Environment Agency (Environment Agency, 2009).

Waterbodies such as the Medlock are identified as HMWBs when physical modifications to the river negatively impact on its quality as a result of the USS. The upstream and downstream pattern in river quality corroborated with the studies of

Guimaràes et al., (2009) who recorded a more diverse invertebrate community at an upstream site which had a vegetation corridor and therefore less pollution from diffuse source runoff. While S1/S2 reflect the lack of direct anthropogenic activity, S4 to S6 are located within the urban area and therefore subject to various impacts such as erosion and modification of substrate, as well industrial and WwTw effluent. Some pollution tolerant taxa including Tubificidae, Lumbricidae, Lumbriculidae,

Chironomidae and Simuliidae were found at the upstream sites. These taxa are most likely to be exhibiting responses to episodic discharges from stormwater outfalls or

CSOs (Grapentine et al. 2004) which are of a sufficient magnitude to select for tolerant benthic taxa. Gammaridae were abundant at the downstream sites. This taxa favours a highly oxygenated river system for reproduction and are present in moderately 146

polluted rivers (Hynes 1970). However, their dominance in certain systems has been linked to their feeding plasticity. Gammaridae fall within the herbivore/shredder guild (Macneil et al. 1997), and hence the presence of organic detritus and other foodstuff within the substrate, plus allochthonous leaf litter and the associated microbial community contribute to an increase in population (Macneil et al. 1997) in the lower Medlock.

SIMPER analysis of the invertebrate assemblage showed the importance of temporal variation as average dissimilarity between winter and summer was 62%.

Oligochaeta, Chironomidae and Simuliidae were abundant in winter compared to summer assemblages. High precipitation during winter increases flow and discharge, increasing run-off and causing instability to the river bed, thereby providing more food for pollution tolerant, fast-growing and fast colonising deposit/suspension- feeding taxa such as Oligochaetes, Chironomidae and Simuliidae (Fonseca and Hart,

1996; Grapentine et al., 2004 and Silveira et al., 2006). During summer Gammaridae increased in abundance, particularly at the downstream sites. Thus, Gammaridae exploit seasonal changes in abundance of specific foods and also are able to rapidly colonise new and variable habitats (Schwartz 1992). Furthermore, the ability of

Gammaridae to exploit a variety of foods is a selective advantage in rapidly changing environments, predating on Asellus as well as exploiting feeding on allochthonous material (Macneil et al. 1997). The Gammarus to Asellus index tend to be more sensitive in organic enriched systems (Whitehurst 1991; Whitehurst & Lindsey 1990) which the Medlock is not and provides further indication of other pollution sources.

The relationship between environmental variables and benthic macroinvertebrate community suggests the impact of discharge, conductivity, altitude, slope, catchment area and nutrients. Indicators of organic pollution, specifically ammonia and BOD did not appear to influence the invertebrate community, probably because both were low due to effective treatment of sewage by the WwTWs and little contribution from CSOs (see Chapter 3). Altitude is a surrogate for river gradient (Roesner & Bledsoe, 2003), slope and catchment area are

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factors which describe the differences between the location and distance of study sites. Increased conductivity at sites downstream between 483 and 693μScmˉ¹ could be related to the catchment geology and the influence of urban streams (Walsh et al.

2005).

Although changes in flow during and after storm events have not been assessed in this study, the LIFE (Lotic Invertebrate-Index Flow Evaluation) index indicates that the river was fast-flowing and therefore could impact on pollution sensitive assemblages (Extence et al. 1999). The influence of flow variability on the benthic invertebrate community will be masked by pollution (e.g. Monk et al., 2006) but the use of the LIFE index here is appropriate in the absence of marked pollution impacts. Other studies have identified the impact of flow on benthic macroinvertebrates including Monk et al. (2006) from a study of 83 catchments in

England and Wales. In this study, various factors including river flow is shown to be a valuable predictor of the instream physical environment and provides a better understanding of river ecosystems than water quality alone (Poff et al. 1997).

PCA which was used to reduce the large number of variables to few parameters suggested that apart from water quality variables, substrate type was one of the factors which structured macroinvertebrate assemblage in this study. This result corroborate other studies that state the suitability of a substrate as a primary factor governing colonisation by benthic macroinvertebrates (Hynes 1970; Silveira et al. 2006). The instability of river substrates is unsuitable for the colonisation of benthic macroinvertebrates as it reduces both diversity and density. Hynes, (1970) and McCulloch, (1986) have reported reduced densities and diversity in sandy and heavily silted streams such that occurs during periods of low flow in the Medlock.

Comparison of BMWP ASPT/NTAXA and the WFD’s WHPTASPT/NTAXA index showed no significant difference in output. However, the similarity of both systems provides an advantage as BMWP serves the basis for communication to non- river ecologists and also for routine monitoring (Armitage et al. 1983; Hawkes 1997) while the WHPT could be used for more technical communication and reports where 148

a greater degree of precision is required. In addition, the similarity allows comparison with the historical data-sets. However, WHPT could provide a better indicator of subtle changes in macroinvertebrate abundance or community structure than the BMWP which may indicate environmental stress (Environment Agency

2015b).

This study suggests that a comparison of observed with reference, pristine conditions determined by RICT for WFD standards is inadequate to classify an urban river. This is because RICT does not take into account direct effects of flow and discharge, plus an indirect effects on substrate arising from the ‘flashy’ nature of urbanised catchments. This study has shown that the quality of urban rivers such as the Medlock is strongly influenced by variability in the magnitude of discharge and frequency of flow and requires an integrated classification tool that accounts for change in flow and discharge. Reducing the magnitude of changes in discharge and flow requires the cooperation of the various stakeholders to facilitate changes to land- use and effluent management to moderate the flow regime, including the application of both hard and soft engineering solutions.

5.6 Conclusion The various multivariate tools and biotic metrics used in this study contributed to show the physico-chemical variables which influenced the River

Medlock. This river is subject to variability in the magnitude of discharge and frequency of flow. The quality of the Medlock, though of good water quality except for PO4-P, is however classed as moderately polluted on the basis of biotic indices and therefore cannot comply with the EU WFD which requires both water chemistry and ecology to achieve “good ecological status”. These changes are due to direct and indirect effects of discharge. Thus, the Medlock, like other urban rivers within the EU is influenced by the integration of factors collectively termed the urban stream syndrome as it is influenced by natural, semi-natural and anthropogenic factors.

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Acknowledgements I am grateful for financial support from The National Open University of Nigeria.

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Barber, D. (2014). Ecological sampling to assess the impacts of metal pollution on Upper River Swale tributaries Gunnerside Gill and Barney Beck. United Kingdom.

Beavan, L., Sadler, J., & Pinder, C. (2001). The invertebrate fauna of a physically modified urban river. Hydrobiologia, 445(1999), 97–108. http://doi.org/10.1023/A:1017584105641

Brittain, J. E., & Eikeland, T. J. (1988). Invertebrate drift - A review. Hydrobiologia, 166(1), 77–93.

Cao, Y., Bark, A. W., & Williams, W. P. (1996). Measuring the responses of macroinvertebrate communities to water pollution: a comparison of multivariate approaches, biotic and diversity indices. Hydrobiologia, 341(1), 1– 19. http://doi.org/10.1007/bf00012298

Chutter, F. M. (1969). The effects of silt and sand on the invertebrate fauna of streams and rivers. Hydrobiologia, 34(1), 57–76.

Clarke, K. R., & Warwick, R. M. (2001). Change in marine communities: An approach to statistical analysis and interpretation. Natural Environment Research Council (2nd ed.). Plymouth, UK: PRIMER-E Ltd Plymouth Marine Laboratory.

Council of the European Union. (2000). Water Framework Directive 2000/60/EC - Official Journal of the European Communities.

Czerniawska-Kusza, I. (2004). Use of Artificial Substrates for Sampling Benthic Macroinvertebrates in the Assessment of Water Quality of Large Lowland Rivers. Polish Journal of Environmental Studies, 13(5), 579–584.

Davies, A. (2002). The use and limits of various methods of sampling and interpretation of benthic macro-invertebrates. Journal of Limnology, 60(Supp. 1), 1–6. http://doi.org/10.4081/jlimnol.2001.s1.1

Davy-Bowker, J., Clarke, R., Corbin, T., Vincent, H., Pretty, J., Hawczak, A., … Jones, I. (2008). River Invertebrate Classification Tool PROJECT WFD72C. SNIFFER. United Kingdom.

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quality streams ? what in we urban can expect. Journal (Water Pollution Control Federation), 54(7), 1139–1147.

Earle, S. (2015). Physical Geology (Open Textb). British Columbia, Canada: British Columbia Open Textbook Project.

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Environment Agency. (2015). Walley Hawkes Paisley Trigg (WHPT ) index of river invertebrate quality A brief description of WHPT for river invertebrate assessment in the UK under the Water Framework Directive .

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Extence, C. A., Balbi, D. M., & Chadd, R. P. (1999). River Flow Indexing Using British Benthic Macroinvertebrates : A Framework for Setting Hydroecological Objectives. Regulated Rivers: Research & Management, 15, 543–574.

Grapentine, L., Rochfort, Q., & Marsalek, J. (2004). Benthic responses to wet- weather discharges in urban streams in southern Ontario. Water Quality Research Journal of Canada, 39(4), 374–391.

Guimaràes, R. M., Facure, K. G., Pavanin, L. a., & Jacobucci, G. B. (2009). Water quality characterization of urban streams using benthic macroinvertebrate community metrics. Acta Limnologica Brasiliensia, 21(2), 217–226. Retrieved from http://ablimno.org.br/acta/pdf/v21n2a210211.pdf

Hawkes, H. (1997). Origin and development of the biological monitoring working party score system. Water Research, 32(3), 964–968. http://doi.org/10.1016/S0043- 1354(97)00275-3

Hellawell, J. M. (1986). Biological indicators of freshwater pollution and environmental management (Pollution Monitoring Series). (K. Mellanby, Ed.). London: Elsevier Applied Science Publishers.

Hynes, H. B. N. (1970). The ecology of running waters. Canada: Univ. of Toronto Press.

James, P., Atkinson, S., Barlow, D., Bates, A., Comyn, F., Duddy, M., … Causer, K. (2012). The Irwell Catchment Pilot:The Rivers Return. Warrington, UK: The Environment Agency.

Langford, T. E., & Bray, E. S. (1969). The distribution of Plecoptera and 151

Ephemeroptera in a Lowland Region of Britain (Lincolnshire). Hydrobiologia, 34(2), 243–271. http://doi.org/10.1007/BF00141931

Lindsay, J. B. (2005). The Terrain Analysis System: A tool for hydro-geomorphic applications. Hydrological Processes, 19(5), 1123–1130. http://doi.org/10.1002/hyp.5818

Lytle, D. A., & Poff, N. L. (2004). Adaptation to natural flow regimes. Trends in Ecology & Evolution, 19(2), 94–100. http://doi.org/10.1016/j.tree.2003.10.002

Macan, T. . (1962). Biotic factors in running water. Aquatic Sciences-Research Across Boundaries, 24(2), 386–407. Retrieved from http://www.springerlink.com/index/L14026L7VMN566H1.pdf\npapers2://publ ication/uuid/CF54EF67-7185-493D-95AD-9D58D3E0425B

Macneil, C., Dick, J. T. A., & Elwood, R. W. (1997). The Trophic Ecology of Freshwater Gammarus Spp. (crustacea:amphipoda): Problems and Perspectives Concerning the Functional Feeding Group Concept. Biological Reviews, 72(3), 349–364. http://doi.org/10.1111/j.1469-185X.1997.tb00017.x

McCulloch, D. L. (1986). Benthic macroinvertebrate distributions in the riffle-pool communities of two east Texas streams. Hydrobiologia, 135(1–2), 61–70. http://doi.org/10.1007/BF00006459

Meier, P. G., Penrose, D. L., & Polak, L. (1979). The rate of colonization by macroinvertebrates on artificial substrate samplers. Freshwater Biology, 9(4), 381–392. http://doi.org/10.1111/j.1365-2427.1979.tb01522.x

Mikalsen, T. (1989). Factors influencing the quality of urban streams in Georgia and the implications for stream management. In Georgia Water Resources Conference (pp. 135–138).

Monk, W. A., Wood, P. J., Hannah, D. M., Wilson, D. A., Extence, C. A., & Chadd, R. P. (2006). Flow variability and macroinvertebrate community response within riverine systems. River Research and Applications, 22(5), 595–615. http://doi.org/10.1002/rra.933

Nilsson, C., & Renöfält, B. M. (2008). Linking flow regime and water quality in rivers: A challenge to adaptive catchment management. Ecology and Society, 13(2). http://doi.org/18

Paisley, M. F., Trigg, D. J., & Walley, W. J. (2014). REVISION OF THE BIOLOGICAL MONITORING WORKING PARTY (BMWP) SCORE SYSTEM: DERIVATION OF PRESENT-ONLY AND ABUNDANCE-RELATED SCORES FROM FIELD DATA. River Research and Applications, 7(4), 887–904. http://doi.org/10.1002/rra

Paul, M. J., & Meyer, J. L. (2001). Streams in the urban landscape. Annual Review of Ecology and the Systematics, 32, 333–365. 152

Poff, N. L., Allan, J. D., Bain, M. B., & Karr, J. R. (1997). Natural Flow Regime. BioScience, (December), 769–784. Retrieved from http://www.jstor.org/stable/1313099

Schwartz, S. S. (1992). Benthic predators and zooplanktonic prey: predation by Crangonyx shoemakeri (Crustacea; Amphipoda) on Daphnia obtusa (Crustacea; Cladocera). Hydrobiologia, 237, 25–30. http://doi.org/10.1007/BF00008424

Silveira, M. P., Buss, D. F., Nessimian, J. L., & Baptista, D. F. (2006). Spatial and temporal distribution of benthic macroinvertebrates in a Southeastern Brazilian river. Brazilian Journal of Biology = Revista Brasleira de Biologia, 66(2B), 623–32. http://doi.org/10.1590/S1519-69842006000400006

Tate, C. M., & Heiny, J. S. (1995). The Ordination of Benthic Invertebrate Communities in the South Platte River Basin in Relation to Environmental- Factors. Freshwater Biology, 33(3), 439–454. Retrieved from ://A1995RE17800010

UK Technical Advisory Group. (2008). The Water Framework Directive UK ENVIRONMENTAL STANDARDS AND CONDITIONS (PHASE 1) Final report April 2008 (Vol. 2008).

Walsh, C. J., Fletcher, T. D., & Burns, M. J. (2012). Urban Stormwater Runoff: A New Class of Environmental Flow Problem. PLoS ONE, 7(9). http://doi.org/10.1371/journal.pone.0045814

Walsh, C. J., Fletcher, T. D., & Ladson, A. R. (2005). Stream restoration in urban catchments through redesigning stormwater systems: looking to the catchment to save the stream. Journal of the North American Benthological Society, 24(3), 690– 705. http://doi.org/10.1899/0887-3593(2005)024\[0690:SRIUCT\]2.0.CO;2

Weber, C. I. (1973). Biological field and laboratory methods for measuring the quality of surface waters and effluents. Cincinnati: U.S. Environmental Protection Agency, Nat. Env. Res. Center, Office Res. Develop.,.

Whitehurst, I. T. (1991). The Gammarus: Asellus ratio as an index of organic pollution. Water Research, 25(3), 333–339. http://doi.org/10.1016/0043- 1354(91)90014-H

Whitehurst, I. T., & Lindsey, B. I. (1990). The impact of organic enrichment on the benthic macroinvertebrate communities of a lowland river. Water Research, 24(5), 625–630. http://doi.org/10.1016/0043-1354(90)90195-C

Willemsen, G. D., Gast, H. F., Franken, R. O. G., & Cuppen, J. G. M. (1990). Urban 153

storm water discharges: Effects upon plankton communities of sessile diatoms and macroinvertebrates. Water Science and Technology, 22(10–11), 147–154.

Wright, I. A., Davies, P., Wilks, D., Findlay, S., Taylor, M. P., Creek, G., & Creek, R. (2007). Aquatic macroinvertebrates in urban waterways : comparing ecosystem health in natural reference and urban streams. In 5th Australian Stream Management Conference. Australian rivers: making a difference. Charles Sturt University, Thurgoona, New South Wales. (pp. 467–472).

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Chapter 6 SHORT TERM WATER QUALITY VARIABILITY IN AN URBAN RIVER SUBJECT TO POINT AND DIFFUSE SOURCE POLLUTION

Abstract Fifteen minute in situ conductivity, turbidity, dissolved oxygen, pH, and temperature measurements plus, continuous discharge data were obtained from 1st

August 2014 to 31st October 2014 at the Environment Agency gauging station on the lower River Irwell, Manchester. This study aimed to determine the impact of combined sewer overflows (CSOs) during short duration events on the basis of the water quality variables. The concentration-discharge relationship showed some variables to be lowered (pH, conductivity, PO4-P and NO3-N) while suspended solids increased with discharge. However, during separation of hydrographs, the study revealed increased PO4-P concentration at high discharge. Peaks of suspended solids and PO4-P observed on the hydrographs suggest spills from CSOs while continuous high concentration during the limb recession points to other pollution sources. All the variables conformed to the requirements of the WFD standards at all discharges apart from PO4-P and suspended solids.

Key words: Conductivity, suspended solids, discharge, gauging station, Water

Framework Directive

6.1 Introduction Most river catchments in the United Kingdom are affected by urbanisation

(Lamb et al. 2003) and its forecast that urban areas will house 68% of the total global population by 2050 (United Nations, 2014) and hence will continue to increase in extent. In the UK, urban areas already contain 79% of the total population and this is projected to increase to 86% by 2050.

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Rivers can be classified based on either topology or discharge and water chemistry (Lindsay, et al. 2008). Two categories based on discharge and water chemistry where identified in the River Medlock; from the river’s source to Lumb

Brook 12km downstream and from Lumb Brook, 5 km to the confluence of the River

Irwell. The current ecological status from the source to Lumb Brook is regarded as

“good” based on Environment Agency’s 2015 prediction while from Lumb Brook to the confluence with the River Irwell the Medlock is designated a ‘heavily modified water body’ (HMWB) and classified as “poor potential” for ecological status and considered at risk (Environment Agency 2015a).

Chapters 3 and 4 examined the long term and medium term variability of the physico-chemical parameters on the Medlock. These studies aimed to establish the historic patterns, spatial and temporal conditions of the river and to identify which variable affected the river’s overall quality. The results suggested that river discharge was an important factor which influenced the physico-chemical variables, plus abundance and distribution of benthic macroinvertebrates (Chapter 5). As the physical variables and, with the exception of PO4-P, chemical indicators of water quality indicated “good” chemical quality, it was expected that the Medlock would also be of good ecological status under the WFD. The degraded invertebrate community was ascribed to direct and indirect effects of changes in discharge and flow.

In this chapter, the short term variability in the water quality of the River

Medlock, was examined at the Environment Agency’s gauging station. By this point the river has received discharges from a variety of sources described above, including

WwTWs, CSOs, surface water drainage and agricultural runoff. This study aims to understand the dynamics of physico-chemical variables in the lower Medlock by examining high resolution datasets that allow an examination of extreme events compared to monthly and or biweekly records (Jarvie et al. 1998). As part of this study, the phenomenon of the “first flush”, described as the first part of runoff which is the most polluted during storm events, is examined. The first flush is composed of

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runoff from rainwater, roofs, discharge from separate and combined sewer systems

(Deletic 1998; Lawler et al. 2006; Lee et al. 2002) and will be examined on the basis of the changes in the physico-chemical variables during storm conditions.

A combination of 15-minute continuously monitored pH, conductivity, turbidity, temperature plus spot samples collected for suspended solids, NO3-N and ammonia-N and PO4-P were examined from the 1st of August to 31st October 2014 at the EA’s London Road gauging station. In addition suspended solids concentration at

15 min intervals was estimated from the relationship between sampled suspended solid concentration and the corresponding turbidity record. The gauging station is approximately 2.9km upstream of the river’s confluence with the Irwell and 6km downstream of the main WwTW on the Medlock.

6.1.1 Aims and objectives The overall aim is to identify the key influences of water quality during storm events by examining high resolution datasets over a period of three months and to determine the changes in the concentrations following continuous precipitation and discharge.

The objectives are:

1. to determine the relationship between pollutant concentration and discharge,

2. to examine the contribution of episodic pollution to the concentration and load

of key variables

3. to estimate the nutrient load exported from the river

The hypothesis is that short-term (less than 24 hr) changes in water quality are due to discharges from combined sewer overflows (CSOs) rather than the WwTW.

6.1.2 Site Description The River Medlock is one of five rivers which form part of the River Irwell catchment. The River Medlock rises to the northeast of Manchester (Figure 6-1) and

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drains a largely urbanised catchment of 57.5km². (National Rivers Flow Archive,

2015).

Water quality data were captured at the EA’s river discharge gauging station

(SJ 848975) 2.9 km from the confluence with the River Irwell in Manchester city centre. At the gauging station, the river has a mean annual flow of 0.819m3s-1 (1974-

2013; National Rivers Flow Archive, 2015). Average annual rainfall over the Medlock catchment between 1961 and 1990 was 1033mm (National Rivers Flow Archive,

2015). The gauged station consists of a non-standard short crested weir 8.5m wide with a sloping downstream face. The weir is located in a rectangular concrete channel with vertical walls upstream of a large culvert. The measurement of discharge is by stilling well and float. The maximum gauged level is 0.48m, maximum gauged flow is 6.75m³sˉ¹ and the bankfull stage i.e. the stage at which a river overflows its natural banks and is likely to cause damage is 3.55m (49.92m³s¯¹).

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Figure 6-1: Location of River Medlock showing catchment area and degree of urbanisation and the gauging station (Source, EDINA).

6.2 Methodology and approach

6.2.1 Continuous sampling programme A pre-calibrated YSI 6600 V2 multiparameter water quality sonde (Yellow

Springs Incorporated, Ohio-USA) was installed adjacent to the gauging station from the 1st of August 2014 until the 31st October, 2014, a total of 92 days. pH, temperature, dissolved oxygen, turbidity and conductivity were recorded at 15-minute intervals.

The sonde was calibrated using a turbidity (0NTU distilled water and Xylem

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Analytics 126NTU suspension), pH (Xylem Analytics pH7 and pH10 buffer solutions) and conductivity (Xylem Analytics 1413µScm¯¹) standards following the manufacturer’s calibration procedures. Dissolved oxygen was calibrated using the

“open-cup” calibration method in which a container is filled with small amount of water which is allowed to equilibrate with the surrounding atmospheric conditions, as per the manufacturer’s procedures and calibrated monthly.

Continuous turbidity measurements allows direct access to the dynamics of particulate pollution (Franklin et al. 2001; Hannouche et al. 2011) and can be used as a surrogate variable for the measurement of suspended solids (Métadier & Bertrand-

Krajewski 2011). In this study, suspended solids measured at the laboratory and the corresponding turbidity sample collected from continuous data were plotted to produce a suspended solids-turbidity linear regression equation y = 0.4405x + 4.486; r

= 0.58, n=48 p<0.0001 (see Appendix). Therefore, on the basis of this relationship, the concentration of suspended solids determined was used in the description of result and analysis of this study.

A 15-minute discharge record was calculated from stage records using a rating curve supplied for the gauging station. Precipitation data was obtained from the

Whitworth Meteorological Observatory (SJ 84681 96760) managed by the Centre for

Atmospheric Science, School of Earth and Environmental Sciences, University of

Manchester.

6.2.2 Spot sampling Water samples were collected from the gauging station in order to establish concentration-discharge relationships. Samples were obtained by lowering a bucket into the river from the bridge located immediately upstream of the gauging station.

These samples were then decanted into a one-litre sample container for subsequent analysis back at the laboratory. Water samples were filtered through a 0.45 μm

Whatman G/F membrane filter and stored at 4°C prior to analysis. Samples were

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analysed for NO3-N, PO4-P, ammonia-N and suspended solids as described above in

Chapter 2.

Hydrograph separation were identified based on high correlation-discharge coefficients (Caissie et al. 1996). This graphical method (Dingman, 2002) was also used in this study as it is straightforward and used by many researchers to describe river hydrodynamimcs (Blume et al. 2007).

Data and statistical analysis was carried out using MS Excel 2010 and

GraphPad Prism version 6.

6.3 Results

Table 6-1 provides a summary of the data collected from the river between

01/08/14 and 31/10/14. The relative standard deviation indicates a large temporal variation for all variables except pH and dissolved oxygen. Discharge varied 40-fold from 0.16m³s-1 to 6.91m³s-1; conductivity ranged from 188µScm-1 to 682µScm-1 with a mean of 519µScm-1; turbidity ranged from 1NTU to 1448NTU with a mean of 34NTU and suspended solids (SS, derived from the SS/turbidity relationship) ranged from 4 mgL-¹to 646 mgL-¹ with a mean of 19mgL-¹. PO4-P was present in very high concentration that varied 7-fold from 0.17mgL-¹ to 1.20mgL-¹ and with a mean of

0.49mgL-¹. NO3-N and ammonia-N were present in very low concentrations in the river with ammonia-N having a minimum concentration below the detection limit of

0.01 mgL-¹ and not exceeding 1.1 mgL-¹and NO3-N having a mean concentration of 3.9 mgL-¹. pH measurements showed the river to be near neutral, ranging from 7.3 to 8.3 and a mean of 7.9 while dissolved oxygen ranged from 59% to 106% with a mean of

84% and water temperature from 9°C to 18°C. Maximum precipitation during recorded the period was 7.3mm.

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Table 6-1: Mean and range of physico-chemical variables in the River Medlock at London Road gauging station between August and October 2014. Data is either from 15-minute continuous analysis or spot samples for ammonia-N, NO3-N and PO4-P. Precipitation is from the Whitworth Meteorological Observatory. N = number of samples

Parameter N Minimum Maximum Mean Std. Dev. RSD (%) Precipitation 8825 0.00 7.37 0.02 0.18 736.44 (mm) Flow velocity 8825 2.70 120.00 8.5 8.3 98.32 (ms¯¹) Discharge 8825 0.16 6.91 0.49 0.48 98.30 (m³s¯¹) Temperature 8825 9.3 18.03 13.00 1.8 13.17 (°C) Conductivity 8825 188.00 682.00 519 86.00 16.67 (µScm¯¹) pH 8825 7.30 8.30 7.90 0.16 1.98 Turbidity 8825 1.00 1448.00 34 103 304.14 (NTU) Dissolved oxygen (% 8825 59.00 106.00 84.00 7.80 9.21 saturation) Suspended 8825 0 646.00 19.00 46 241.85 solids (mgL¯¹) Ammonia-N 48 ≤0.01 1.1 0.34 0.24 72.00 (mgL¯¹)

NO3-N (mgL¯¹) 50 1.1 8.40 3.9 2.00 52.41

PO4-P (mgL¯¹) 50 0.17 1.20 0.49 0.25 51.54

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6.3.1 Discharge and Precipitation Table 6-2 presents discharge and corresponding precipitation records during the study and the hyetograph of the relationship is shown on Figure 6-2. The mean discharge at the gauging station measured for the sampling duration was 0.49m3s-1, mean precipitation was 0.02mm, and total precipitation recorded was 209.69mm.

August 2014 was very wet, including numerous days with light rainfall; no rain fell only on 4th and 23rd August 2014. The highest precipitation and peak discharge

(6.91m³s⁻¹) was recorded on the 11th August 2014. As shown on Table 6-2 other high discharges were also recorded in August 2014 while the lowest records were obtained in September 2014 (5th & 23rd September 2014) and from 11th to 13th October

2014. A significant correlation was found between discharge and precipitation

(Pearson correlation, n=8825, p < .01) which, as expected, showed that there was a trend of increasing water volume with increased rainfall.

Table 6-2: Total precipitation and average discharge from 1st August to 31st October 2014.

Sample Total precipitation Average discharge Number of days months (mm)/month (m3s-1)/month Aug-14 31 128.12 0.68 Sep-14 30 17.7 0.28 Oct-14 31 63.87 0.5 Total 92 209.69

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Figure 6-2: Hyetograph taken over 15-minute duration from 1st August 2014 to 31st October 2014.

6.3.2 Correlation of physico-chemical variables with discharge and intercorrelation between variables The relationship between, respectively, conductivity, pH, precipitation and spot samples of NO3-N, PO4-P, ammonia-N and mean continuous discharge over 90 days is shown in Figure 6-3. 15-minute interval measurements are shown for pH, conductivity and suspended solids (Figure 6-3A-C) while discharge readings corresponding to daily spot sample were plotted for PO4-P (r = -0.26, n=50, p>0.05);

NO3-N (r=-0.42, n=50, p<0.05) and ammonia-N(r=0.22, n=48, p>0.05) (Figure 6-3D-F).

Discharge was negatively correlated with all the variables (conductivity r= -0.69, p<0.0001, n=8825; pH r= -0.36, p<0.0001, n=8825) except suspended solids (r = 0.18, p<0.001, n=8825) and ammonia-N which showed a positive relationship. Suspended solids vs discharge plotted in Figure 6-3C was influenced by the river’s desiltation during the sampling period (see hydrograph of Figure 6-6B). While an increase in discharge indicated the dilution of nutrients and dissolved salts, a positive correlation with suspended solids suggested other sources of pollution to the river including road runoff and CSOs. No correlation was found between discharge and ammonia-N which indicates that the concentration of ammonia-N in the river was independent of river discharge.

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Suspended solids (D) PO4-P (E) NO3-N and (F) ammonia-N. Figures A, B and C are based on 15- minute discharge while D, E and F correspond to the dates when spot samples were collected from the river.

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B A

Figure 6-4: Relationship between physico-chemical concentrations measured in the River Medlock at the gauging station between 1st August 2014 to 31st October 2014. (A) Mean daily pH and mean daily conductivity (B) NO3-N and PO4-P

Figure 6-4 shows the strongest correlation to be between NO3-N and PO4-P (r =

0.87) and conductivity and pH (r = 0.43).

6.3.3 Temporal variability Monthly changes in the physico-chemical parameters in the river from 1st

August 2014 to 31st October 2014 are presented in Figure 6-5A-I. All the variables showed a significant (p<0.05) difference with month except ammonia-N. As expected, there was a gradual decline in temperature from August to October. There was consistently high percentage saturation of dissolved oxygen in the river. Very high concentration of suspended solids (>25mgL¯¹) were recorded when compared to the standard annual mean of 25 mgL¯¹ under the EU Freshwater Fish Directive (DEFRA,

2010). PO4-P concentration classified the river as “poor” under the WFD over more than 80% of the sampling period as the concentration exceeded the 0.12mgL¯¹ standard. Concentrations indicative of “moderate” pollution of between 0.12mgLˉ¹ and 0.25mgL¯¹ were recorded, mainly in October. Mean concentration of ammonia-N complied with the WFD standard of 0.6 mgL¯¹; however a few samples in August had moderately high ammonia-N concentrations with a range of 0.79 to 1.06mgL¯¹.

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15-minute continuous data collection or spot sampling between 1st August 2014 and 31st October

2014. (A) Discharge, (B) Conductivity, (C) pH, (D) Temperature, (E) Dissolved oxygen, (F)

Suspended solids, (G) Ammonia-N, (H) NO3-N, (I) PO4-P.

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6.3.4 Chemical concentration vs discharge On the basis of the correlation found between the chemical parameters and discharge, hydrographs were plotted for conductivity and suspended solids by using the continuous measurement discharge records (Figure 6-6) and spot samples for

PO4-P and NO3-N for the sample period (Figure 6-7).

Conductivity declined following high discharge as highlighted by the black boxes on 1st August, 29th August 2014, 6th and 7th October (Figure 6-6A) indicating dilution by run-off that presumably contains large quantities of low conductivity rainwater. Suspended solids concentration was shown to be very high on 14 August

2014, 22- 28 September and 25 -29 September 2014 (Figure 6-6B). While the record suggests sediment influx from the surrounding catchment which has been washed into the river by rainfall events prior to 14th August, some days after the high precipitation, the river still recorded high concentration of suspended solids.

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Figure 6-6: Hydrographs of (A) Conductivity; (B) Suspended solids over 15-minute intervals from 1st August and October 2014.

The time series for continuous discharge and nutrients (Figure 6-7) showed high nutrient concentration during periods of low discharge; a pattern which suggests the effect of the discharge from the WwTWs. Occasionally, such as 10th

August 2014, a very high PO4-P level was recorded (0.73mgL¯¹) with increased discharge, 6.82m³s¯¹ which suggests contributions from episodic point source, probably CSOs, following storm events. The 1st of August provided a good example of the variability in nutrient concentration at high discharge. At 11:00 a discharge of

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0.19m3s-1, had a corresponding increase concentration of PO4-P of 0.82mgL¯¹; and NO3-

N of 6.43mgL¯¹; at 15:45, an increase in discharge to 0.44m³s-¹ resulted in 1.12mgL-1 of

PO4-P and 8.38mgL¯¹ NO3-N at 20:00hr. A further increase in discharge to 1.10m³s-¹ however resulted in fall in the PO4-P concentration to 0.68mgL⁻¹ and NO3-N to

3.89mgL¯¹. This implied that at the initial periods, where high PO4-P concentration was recorded, the impact of episodic discharges (“first flush”) triggered by increases in discharge was followed by dilution with continuous discharge. These results indicate that both WwTW and CSOs contribute to PO4-P concentration in the river although the latter only during initial periods of high rainfall. Neither WwTWs nor

CSO are significant sources of NO3-N.

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Figure 6-7: Continuous time series of discharge in the River Medlock from 1st August 2014 to 31st

October 2014 and concentrations of (A) PO4-P and (B) NO3-N from spot sampling during low and high flows.

Peak discharge events were selected for further analysis using hydrograph separation to study the impact of storm events on chemical concentrations as seen in other urban systems (e.g. Pilgrim et al. 1979). On the basis of peak discharge identified in earlier hydrographs, the 10th August 2014 which was one of the periods

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with the highest river discharge at 6.82m³s-¹ was used for hydrograph separation in order to determine the effect of rainfall on conductivity, pH and suspended solids

(Figure 6-8A-C). Conductivity decreased with increasing discharge indicating some dilution. Continued precipitation led to a further increase in discharge from 1.53m³s-¹ at 2100hr to 6.82m³s-¹ at 2315hr followed by an initial slight increase in conductivity and, with continuous precipitation followed a progressive decrease in conductivity levels. Increased precipitation over two hours increased discharge retention leading to decreased conductivity. pH followed a similar pattern to conductivity but the change was small with a fall of around 0.4 for one hour. Suspended solids correlated very strongly with discharge (Figure 6-8C), (r=0.928, p < .05, Two-tailed t-test) with both increasing proportionately.

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Figure 6-8:Hydrograph separation of high discharge on the 10th August 2014 using 15-minute continuous data for (A) conductivity; (B) pH; (C) suspended solids.

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Figure 6-9 and Figure 6-10 below show the response of the river during two rainfall events; long-duration at 27mm for 11 hours 15 minutes on the 1st August 2014

(Figure 6-9) and short-duration 15mm rainfall for 1hour 30 minutes on the 8th of

August 2014 (Figure 6-10). Total precipitation recorded on the 1st August 2014 when it rained continuously for 11 hours 15 minutes between 1045hr and 2200hr, was

27mm. There was no recorded change in discharge and conductivity (Figure 6-9A) for five hours when rainfall rate increased from 0.1mm to 6.76mm by 1530hr.

Conductivity levels decreased after 30 minutes by 16% from 674µScm¯¹ to 562µScm¯¹.

A decrease in precipitation and increasing discharge led to lower conductivity, falling to 364 µScm¯¹ (i.e. 46% less than the initial value) at 2015hr followed by an increase to

518µScm¯¹ at 2100hr and then a decrease to 362 µScm¯¹ at a discharge of 1.58m³sˉ¹ when the rain stopped four hours later. The converse was the case with suspended solids (Figure 6-9B) which increased with increasing precipitation and discharge, from 8mgL- ¹ to 46mgL¯¹ at 2145hr. As earlier observed, there is a significant (r²= 0.939, p < .001) correlation between suspended solids and discharge.

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Figure 6-9: Hydrograph separation for the long duration rainfall (27mm) event on 1st August 2014 using 15-minute continuous reading for (A) Conductivity (B) Suspended solids with discharge data and total precipitation. The ellipse highlights the decrease in in conductivity levels with the rising limb of the hydrograph and increase in suspended solids concentration.

A slightly different pattern was found for the shorter duration rainfall event at

15mm (Figure 6-10 A and B) on the 8th August 2014 when it rained from 1345hr to

1515hr. As with the long duration rainfall event, conductivity declined when discharge started to rise and precipitation was at its peak at 7.4mm/day. However, the recovery time was faster due to the short duration of the rainfall event and decline in discharge. A higher suspended solids concentration was reached to

92mgL¯¹ but declined after six hours (2315hr) to 17mgL¯¹ after rainfall ceased. Both conductivity and suspended solids concentration recovered to their pre-rainfall concentrations faster than during the continuous rainfall event indicating duration of rainfall has a strong influence on the concentration and that larger events are more 175

likely to have the “first flush” event. The peak concentration associated with the rising limb of the hydrograph for suspended solids (Figure 6-9B and Figure 6-10B) suggests mobilisation of sediment and/or first flush of particulates from combined sewer systems (CSS) via CSOs into the river. This was followed by storm water dilution and the consequent recovery to base flow conditions.

Figure 6-10: Hydrograph separation of 8th August 2014 total rainfall event at 15mm using 15min continuous reading for (A) Conductivity (B) Suspended solids with discharge and total precipitation *the ellipse indicates increase in concentration with the rising limb of the hydrograph.

The ellipses on Figure 6-10 A and B highlight the effect of discharge on conductivity and suspended solids.

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6.3.5 PO4-P Load

The PO4-P load estimated at the gauging station showed that August 2014 had the highest load record 18kgPhaˉ¹yearˉ¹ as shown on Table 6-3 due to increased precipitation and hence discharge from episodic point sources such as CSOs.

However, daily load showed the values were less than 0.1kgPhaˉ¹dayˉ¹.

Table 6-3: Total precipitation and PO4-P load from the river Medlock from 1st August to 31st October 2014.

Sample months Total PO4-P load PO4-P load Precipitation/month (kgha ˉ¹yearˉ¹) (kgha ˉ¹ dayˉ¹) (mm)

Aug-14 128.12 18.25 0.05

Sep-14 17.70 1.95 0.01

Oct-14 63.87 2.21 0.01

In summary, all the variables analysed in this study showed dilution with increased discharge apart from ammonia-N and suspended solids. The concentration of suspended solids recorded during the period following increased precipitation could be as high as 600mgLˉ¹ indicating the effect of “first flush”. This was followed by dilution with continuous discharge. Hydrograph separation served to show the trend of selected variables over a sample time and the changes that occurred during the period of sampling. With reference to compliance, the mean PO4-P was higher than WFD requirement of 0.12mgLˉ¹ and during the period of study, the standard was not achieved even when there was no rainfall/high discharge which indicates the continuous influence of the treatment works. Evidence of “first flush” episodic phenomena was observed for PO4-P in August when a high precipitation/discharge corresponded to a very high concentration.

During long duration rainfall, the recovery time taken for concentration of the measured water quality parameters to return to their pre-storm levels was slow, while the reverse is the case for short duration rainfall event. Apart from ammonia-N 177

which showed no significant change, all variables measured were significantly different between the sample months. PO4-P load estimated during the sampling period indicated that August 2014 was impacted by episodic pollution due to high rainfall events. Other sampled months had <3kgPhaˉ¹yearˉ¹, which is similar to the estimated load during the fortnightly sampling period (Chapter 4).

6.4 Discussion The short-term temporal dynamics of the River Medlock was assessed from continuous monitoring of key water quality parameters. This included hydrograph separation for conductivity, pH and suspended solids following examination of the concentration-discharge relationships which indicated the variables to be separated by hydrographs. This pattern was used by Caissie et al., (1996) to determine the relative contribution of groundwater (pre-event) flow to total flow (or event water) during storm events of different magnitude.

The hydrographs were ‘flashy’, rising steeply with minor attenuation especially during high flows. Peaks in discharge which correlated positively with rainfall were observed for suspended solids and occasionally for PO4-P. Decreased conductivity and pH indicate dilution with increased discharge. High conductivity during autumn and winter could be linked to the runoff of de-icing salts into the river. While NO3 and PO4-P responded to stream flow by dilution, ammonia-N and suspended solids did not follow the same pattern. This implies that river concentrations during the high precipitation “first flush” event could be linked to the duration, frequency and magnitude of spills which could then impact on urban river quality.

Turbidity was used as a surrogate variable for suspended solids concentration.

(Bilotta & Brazier 2008). The high concentration of suspended solids observed in the

Medlock in October 2014 (6th, 22nd and 27th) was not “first flush” event but was attributed to the desilting of the river by the Environment Agency as part of flood defence and control activity (EA personal communication, 2015) and exacerbated by 178

high rainfall as was the case in August. Also, high concentrations recorded in the river some days after the rainfall could be linked to “late” arrival of suspended materials in the storm (i.e. reverse first flush) indicating a departure from “first flush models”(Lee et al. 2002). This delay indicates that CSOs were not the major contributors to the changes observed in the river at short duration high flows.

Increase in rainfall increases the hydraulic gradient and also saturates the soil (Blume et al. 2007). Thus, elevated suspended solids during high discharge could indicate upstream erosion (Chebbo et al, 2001) as increased flow from the upper sub- catchments resulted in the sediments being transported downstream to the gauging station as shown by increased concentration of suspended solids from 22mgLˉ¹ at

2145hr to 167mgLˉ¹ at 2330hr. Such high suspended solid concentrations are likely to adversely affect the benthic macroinvertebrates (Beck et al. 2004; Newcombe &

Macdonald 1991; Crabtree 1989) and hence contribute to the degraded community in the River Medlock (Chapters 3 and 5).

Concentrations of both nutrients were highest in August reflecting the seasonality in precipitation and fertiliser application. Agricultural fertilisers are applied to the soils during August in the UK (Farming and Countryside Education

(FACE) 2007) and the high rainfall in August (total precipitation: 128mm) will transport the fertilisers to the Medlock. This is exacerbated by crop harvesting and planting of winter crops as land is left bare and ploughed; such activities also provide a potential sediment source to the Medlock. Apart from August 2014, PO4-P load estimated for the three month period had similar output (<3kgPhaˉ¹yearˉ¹) for the low resolution and monthly spatial datasets (Chapters 4 and 5 respectively). The results there also showed that a reduction of effluent PO4-P from the WwTWs would reduce the concentration and load. The various datasets have revealed that the single

WwTWs is the largest source of PO4-P and suspected by the water company (United

Utilities, personal communication, , 2014).

Increased PO4-P and suspended solids concentration identified in this study showed that the river is subject to episodic conditions which influence its quality. 179

Although maximum concentrations were recorded at the beginning of the storm events, the recovery time of the river varied with the duration and volume of rainfall/runoff. Thus, CSO effects are temporal while other pollution sources including WwTW, diffuse runoff are continuous and these are integrated in the quality of the Medlock.

High dissolved oxygen levels recorded during the sampling period showed that the effect of organic pollution as a result of “first flush” is limited, in part due to the high degree of re-aeration resulting from the turbulent flow. pH was near-neutral and also correlated with conductivity and both variables were diluted with increased rainfall. This further implies that the “first flush” activity was rare. This study has shown that the urban River Medlock especially the part classified as a “highly modified water body” is subject to urban stress which is exacerbated by increased precipitation and discharge from point and diffuse sources. Such stormwater run-off combined with hydraulically efficient drainage is a key reason why the Medlock does not display “good ecological potential” as required by the WFD.

6.5 Conclusion High resolution datasets provided information on the short term variability of physico-chemical variables at the River Medlock. The potential for “first flush” events markedly influencing water quality was shown by PO4-P and suspended solids. The duration of rainfall events can have a strong influence on the river’s water quality due to rapid dilution of first-flush pollutants. Concentrations increased during large storm events and gradually returned back to pre-event concentrations.

Simultaneously, other variables were diluted with rainfall-induced increases in discharge. The results show that the activities of CSOs were short-lived and that the

Medlock was subject to other point and diffuse sources which may affect its quality.

As a highly modified water body, the River Medlock is subject to urban stress due to the hydraulically efficient drainage; therefore at high discharge it is influenced by

180

water quantity as well as quality. This study does not support the hypothesis that short-term water quality is solely due to discharges from CSOs.

Acknowledgement:

I am grateful for financial support from The National Open University of Nigeria. Thanks to Cascade Consulting (now Renovo), Manchester for the use of their YSi meter and data collection at the Medlock, and Dr Gareth Martins of Cascade for help and advice with the installation.

6.6 References Beck, B., Old, G. H., Leeks, G. J. L., Packman, J. C., Smith, B. P. G., Lewis, S., & Hewitt, E. J. (2004). Physical and chemical extremes of the urban river environment :, II, 317–325.

Bilotta, G. S., & Brazier, R. E. (2008). Understanding the influence of suspended solids on water quality and aquatic biota. Water Research, 42, 2849–2861. http://doi.org/10.1016/j.watres.2008.03.018

Blume, T., Zehe, E., & Bronstert, A. (2007). Rainfall – runoff response, event-based runoff coefficients and hydrograph separation. Hydrological Sciences Journal, 52(5), 843–862. http://doi.org/10.1623/hysj.52.5.843

Caissie, D., Pollock, T. L., & Cunjak, R. a. (1996). Variation in stream water chemistry and hydrograph separation in a small drainage basin. Journal of Hydrology, 178, 137–157. http://doi.org/10.1016/0022-1694(95)02806-4

Crabtree, R. W. (1989). Sediments in Sewers. Water and Environment Journal, 3(6), 569–578. Retrieved from http://dx.doi.org/10.1111/j.1747-6593.1989.tb01437.x

Defra. (2010). Freshwater fish directive. Retrieved from http://archive.defra.gov.uk/environment/quality/water/waterquality/fwfish/ind ex.htm

Deletic, A. (1998). The first flush load of urban surface runoff. Water Research, 32(8), 2462–2470. http://doi.org/10.1016/S0043-1354(97)00470-3

Environment Agency. (2015). Catchment Restoration Fund: Environment Agency Final Annual Report 2014-2015.

Farming and Countryside Education (FACE). (2007). Discovering Farming in Britain. United Kingdom.

Franklin, D. H., Steiner, J. L., & Wheeler, G. (2001). Comparison of different methods of measuring turbidity for estimation of total suspended sediments. In A. Hatcher, K.J Institute of Ecology, University of Georgia (Ed.), Georgia Warer 181

Resources Conference. Georgia.

Hannouche, A., Chebbo, G., Ruban, G., Tassin, B., Lemaire, B. J., & Joannis, C. (2011). Relationship between turbidity and total suspended solids concentration within a combined sewer system. Water Science and Technology, 64(12), 2445–2452.

Lamb, R., Zaidman, M. D., Archer, D. R., Marsh, T. J., & Lees, M. L. (2003). River Gauging Station Data Quality Classification (GSDQ) - R&D Technical Report W6- 058/TR.

Lawler, D. M., Petts, G. E., Foster, I. D. L., & Harper, S. (2006). Turbidity dynamics during spring storm events in an urban headwater river system: the Upper Tame, West Midlands, UK. The Science of the Total Environment, 360(1–3), 109– 26. http://doi.org/10.1016/j.scitotenv.2005.08.032

Lee, J. H., Bang, K. W., Ketchum, L. H., Choe, J. S., & Yu, M. J. (2002). First flush analysis of urban storm runoff. The Science of the Total Environment, 293(1–3), 163–175. http://doi.org/10.1016/S0048-9697(02)00006-2

Lindsay, J. B., Rothwell, J. J., & Davies, H. (2008). Mapping outlet points used for watershed delineation onto DEM-derived stream networks. Water Resources Research, 44(8).

Métadier, M., & Bertrand-Krajewski, J. L. (2011). From mess to mass: A methodology for calculating storm event pollutant loads with their uncertainties, from continuous raw data time series. Water Science and Technology, 63(3), 369–376.

Nations, U. (2014). World Urbanization Prospects, the 2014 Revision. http://doi.org/10.4054/DemRes.2005.12.9

Newcombe, C. P., & Macdonald, D. D. (1991). Effects of Suspended Sediments on Aquatic Ecosystems. North American Journal of Fisheries Management, 11(1), 72– 82.

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Chapter 7 SUMMARY AND CONCLUSIONS

The first research chapter (three) provided an overview of the state of the

Medlock over the previous decade with the aid of EA long-term datasets from three sites on the lower reaches. Interrogation of such data provides an indication of trends in water quality and allows the effectiveness of management strategies and policy changes to be assessed. Although some data was not available for two sites, analysis of the results suggests the pollution was largely non-sewage related as BOD and ammonia-N were low. Although PO4-P declined with time the concentration was still higher than the WFD standard. This data reflects the current policy of reducing BOD, ammonia and suspended solids from WwTWs, but not always PO4-P due to the cost of tertiary treatment to remove this compound (Mainstone et al., 2000; Mainstone &

Parr, 2002). Occasionally, high suspended solids recorded over the study period indicate the impact of high precipitation events. This chapter identifies the WwTWs as a key influence on river quality, specifically PO4-P.

In chapter four, the contribution of WwTW, CSOs and other sources to the high PO4-P loading was investigated with the use of the long-term EA data plus fortnightly data from a larger number of sites to improve the spatial resolution and data from a high temporal resolution study at a single site. The results suggested that the WwTW was a major contributor to the PO4-P load with an average of 92% from this source. The remainder of the contribution was associated with CSOs and other diffuse sources and which varied with discharge. Non-WwTWs sources of PO4-P were greatest during storm events during initial high discharge. It also showed that reduction of PO4-P from the WwTW would reduce the river load considerably. A comparison of total phosphorus load per unit area of catchment (<3kgTPha⁻¹yr⁻¹) estimated in this study was commonly around a third of total phosphorus load obtained in the literature for other urbanised areas.

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Chapter five demonstrates the usefulness of multivariate techniques and interactions within and between samples/variables and, the potential of biotic indices in determining the benthic status of urban rivers such as the Medlock. The study revealed that the degraded benthic community in the Medlock was not primarily due to poor water quality but as a direct and indirect result of water quantity. With the exception of PO4-P, water quality indicators including DO, ammonia and BOD conformed to the requirements of the WFD, including at high discharge and are hence unlikely to adversely impact on the invertebrate community. Tolerance to disturbance and suspended solids plus feeding habits and ability to rapidly colonise a habitat were the main factors determining survival. Therefore, taxa such as

Oligochaeta and Chironomidae that are resistant to the instability caused by changing flow conditions in the river dominate. Gammaridae were also common at the polluted sites, particularly during the summer, as they are also able to take advantage of the plentiful supply of allochthonous material. Both the BMWP and the

WHPT are equally effective at detecting change in the invertebrate community despite the latter also including a measure of abundance that may have been expected to have resulted in greater discrimination between sites.

In chapter six, high resolution sampling confirmed that Medlock water quality was influenced not only by CSOs during high discharge events but was also subject to other point and diffuse sources. Increased concentration of suspended solids and

PO4-P at peak discharge indicated the effect of water quantity on the river’s quality.

This part of the study has brought further clarity to an understanding of the behaviour of the River Medlock especially as it pertains to the unstable conditions that degrade the benthic macroinvertebrate community as shown in Chapter 5.

The individual chapters have highlighted the challenges to the River Medlock, and hence other urban rivers in achieving the required WFD standards. The non- achievement of this standard is linked to the contribution of both continuous, episodic point and possibly diffuse sources that influence both the quantity and

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quality of the receiving water. This study also showed that effective sampling in time and space is necessary to describe patterns and processes. While the long term datasets provided an overview, the fortnightly sampling provided a more focussed and spatially coherent sampling of both water quality and ecology. This enabled the characterisation of river quality and the benthic invertebrate community. The high resolution sampling quantified the scale of the changes during peak discharge and re- emphasized the impact of diffuse pollution and effects of the urban stream syndrome.

Like the Medlock, PO4-P pollution has been a major problem in Europe and the US, its management to date has focussed on controlling its removal from municipal wastewater. Almost half of the rivers and three-quarters of lakes in

England suffer from eutrophication as a result of elevated phosphorus concentration

(Davey & VanLeirde 2015). Substrate instability is a further problem highlighted by this study resulting in peaks of suspended solids during storms which exceeded the

EU Freshwater Fisheries Directive of ≤25mgLˉ¹ for 9% of the storm duration. Finally, the Medlock is characterised by a ‘flashy’ hydrograph characteristic of urban rivers

(e.g. Paul & Meyer 2001). This study therefore does not support the hypotheses that

CSOs are usually the key contributor to the poor water quality and reduction in the benthic macroinvertebrate community in the river.

The term ‘‘urban stream syndrome’’ as described in Chapter 5 was coined by

(Meyer et al. 2005) to describe the long-term ecological degradation of rivers and streams draining urban catchments. Key symptoms of the urban stream syndrome are a flashy hydrograph, high nutrients and other contaminants, altered channel morphology due to re-engineering, and reduced biodiversity and increased dominance. The mechanisms driving the urban stream syndrome are numerous and interactive, but result from a few major sources, specifically storm water runoff but also CSOs and WwTWs plus the legacy of pollution from previous activities within the catchment (Meyer et al. 2005;Walsh et al. 2005). On the basis of the above definition and criteria, there is no doubt that the Medlock suffers from the urban

185

stream syndrome. Because of the urban stream syndrome, “Good” water quality as defined by the WFD will not result in “Good Ecological Status” for the River

Medlock.

In order to effectively manage and control pollution, it would be important to reduce the frequency and intensity of river discharge through hard and soft engineering which will promote biodiversity at the Medlock. Some measures to reduce water quantity is the construction of sustainable urban drainage systems

(SuDS)(Maltby 2012). SuDS are systems which mimic natural systems such as

‘raingardens’ to drain surface water and release it slowly back into the environment.

Such approaches aim to address urban flooding and sewage overflow while promoting urban greening. The sustainability of urban rivers would be enhanced by installing permeable pavements which increase infiltration and, also to involve and link various stakeholders who will ensure continuity of green infrastructural improvements (Royal Geographical Society (with IBG) 2012). Furthermore, the use of wetland vegetation could be explored to reduce pollutants such as phosphorus in the less urban areas of the Medlock catchment where land is available such as Clayton

Vale. Such wetlands will also reduce the risk of flooding down-stream. Such a scheme is being installed on the lower Irwell where a five-hectare wetland will also increase biodiversity by containing marsh, reed-beds, gravel islands and ponds

(SalfordOnline, 2015). The water company, United Utilities is constructing an underground storage tank which will hold excess storm water of around 17,800 m3sˉ¹ during storm conditions. This storm tank is expected to reduce the frequency of CSO releases to the Irwell from Davyhulme WwTW (United Utilities http://www.unitedutilities.com/2601.aspx). The Medlock would benefit from a similar installation to reduce the frequency of CSO discharges and hence levels of suspended solids. However, this would not reduce total phosphorus load in the absence of subsequent treatment. PO4-P concentration in the River Medlock can be reduced by stripping from the WwTW effluent as has been explored elsewhere (Neal et al. 2005) and to improve agricultural practices upstream of the river. 186

Further work

 To investigate the effect of sediment traps on the river ecology

 To investigate the impact of extreme conditions in relation to climate change

(increased/ reduced rainfall) and the resilience of urban streams.

7.1 Conclusion The overall aim of the thesis is to examine the impact of point-source pollution from WwTWs and CSOs on the water quality. While the WwTW was shown to be a major source of PO4-P pollution, the effect of other sources including CSOs, runoff and high precipitation were revealed. The benthic invertebrate community are influenced mainly by the direct and indirect effects water quantity, specifically the highly variable (‘flashy’) hydrograph rather than river quality. The introduction of storm water control measures would mitigate pollution entering the River Medlock.

This could include the construction of storm water purification tanks at the Medlock for the control of extreme flood events, thereby reducing flow, suspended solids and nutrients.

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Appendix

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Table i. Concentration of suspended solids measured at the river at the EA’s gauging station and corresponding turbidity levels

Date Time (hr:mm:ss) Turbidity (NTU) SS (mgL¯¹) 01-08-14 11:15:52 2.2 4.00 01-08-14 16:00:52 24.9 3.00 01-08-14 20:15:52 52.4 48.00 02-08-14 12:15:52 18.4 9.67 02-08-14 14:00:52 7.9 6.67 02-08-14 18:00:52 6.7 14.67 03-08-14 17:45:52 1.8 5.67 05-08-14 12:30:52 2.2 5.00 06-08-14 13:00:52 3.1 0.00 08-08-14 15:15:52 136.9 50.00 08-08-14 15:30:52 108.2 41.33 08-08-14 17:30:52 148.6 237.33 10-08-14 09:45:52 11.8 4.33 10-08-14 12:00:52 103.1 107.33 10-08-14 16:45:53 93.3 33.00 10-08-14 17:30:52 72.5 50.33 11-08-14 16:00:53 23 6.33 11-08-14 18:45:52 34.4 13.33 12-08-14 11:30:52 11.4 9.00 12-08-14 16:30:52 8.3 6.00 13-08-14 13:00:52 11.3 7.33 14-08-14 14:00:52 238.7 50.00 14-08-14 18:00:52 80.6 87.67 17-08-14 14:00:53 6.4 6.33 17-08-14 18:30:53 6.4 4.33 18-08-14 09:45:53 8 4.00 19-08-14 10:00:53 5.1 1.00 27-08-14 19:45:53 185.3 8.00 28-08-14 12:00:52 2.1 0.00 04-09-14 12:45:53 2.5 0.00 09-09-14 18:00:53 2.6 9.00 10-09-14 10:30:53 4.7 11.33 15-09-14 17:30:53 12.6 11.33 16-09-14 13:00:53 27.7 7.67 19-09-14 10:15:52 4.5 0.00 20-09-14 16:00:53 4.4 23-09-14 17:15:53 28 0.00 23-09-14 20:00:53 16.1 1.67 24-09-14 11:30:53 14.7 3.00 26-09-14 14:00:53 48.6 2.33 29-09-14 12:00:53 2.5 0.00 30-09-14 09:30:53 3 0.00 30-09-14 16:00:53 2.8 25-10-14 18:15:54 3 0.00 26-10-14 15:00:54 3.8 0.00 27-10-14 18:45:54 7.2 0.00 28-10-14 15:00:54 40.4 9.67 29-10-14 17:30:54 8.9 5.33 30-10-14 12:00:54 18.1 11.00 31-10-14 16:15:54 3.1 53.60

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