SILTFLUX Literature Review - EPA

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Report No. 176

SILTFLUX Literature Review Authors: D. Lawler, A. Rymszewicz, L. Conroy, J. O’Sullivan, M. Bruen, J. Turner, M. Kelly-Quinn

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EPA RESEARCH PROGRAMME 2014–2020

SILTFLUX Literature Review (2010-W-LS-4) EPA Research Report Prepared for the Environmental Protection Agency by University College Dublin and Coventry University Authors: Damian Lawler, Anna Rymszewicz, Liz Conroy, John O’Sullivan, Michael Bruen, Jonathan Turner and Mary Kelly-Quinn ENVIRONMENTAL PROTECTION AGENCY An Ghníomhaireacht um Chaomhnú Comhshaoil PO Box 3000, Johnstown Castle, Co. Wexford, Ireland Telephone: +353 53 916 0600  Fax: +353 53 916 0699 Email: [email protected]  Website: www.epa.ie

© Environmental Protection Agency 2017

ACKNOWLEDGEMENTS

This report is published as part of the EPA Research Programme 2014–2020. The programme is financed by the Irish Government and administered by the Environmental Protection Agency, which has the statutory function of co-ordinating and promoting environmental research. The project team would like to acknowledge the very valuable input from the project steering committee, Professor Des Walling, Professor John Quinton, Professor Steve Ormerod, Dr Martin McGarrigle, Catherine Bradley, Colin Byrne, Marie Archbold, Wayne Trodd, Donal Daly and Alice Wemaere. In addition, the expert advice of Dr Martin O’Grady in the early stages of the project was invaluable.

DISCLAIMER

Although every effort has been made to ensure the accuracy of the material contained in this publication, complete accuracy cannot be guaranteed. Neither the Environmental Protection Agency nor the authors accept any responsibility whatsoever for loss or damage occasioned, or claimed to have been occasioned, in part or in full, as a consequence of any person acting, or refraining from acting, as a result of a matter contained in this publication. All or part of this publication may be reproduced without further permission, provided the source is acknowledged. The EPA Research Programme addresses the need for research in Ireland to inform policymakers and other stakeholders on a range of questions in relation to environmental protection. These reports are intended as contributions to the necessary debate on the protection of the environment.

EPA RESEARCH PROGRAMME 2014–2020 Published by the Environmental Protection Agency, Ireland ISBN: 978-1-84095-645-0

February 2017

Price: Free

Online version ii

Project Partners

Professor Michael Bruen School of Civil Engineering, UCD Dooge Centre for Water Resources Research and UCD Earth Institute University College Dublin Belfield Dublin 4 Ireland Tel: +353 1 716 3212 Email: [email protected]

Dr Jonathan Turner School of Geography and UCD Earth Institute University College Dublin Belfield Dublin 4 Ireland Tel. +353 1 716 8175 Email: [email protected] Professor Damian Lawler Centre for Agroecology, Water and Resilience James Starley Building Coventry University Coventry CV1 5FB Tel: +44 2477 651674 Email: [email protected]

Dr John O’Sullivan School of Civil Engineering, UCD Dooge Centre for Water Resources Research and UCD Earth Institute University College Dublin Belfield Dublin 4 Ireland Tel: +353 1 716 3213 Email: [email protected]

Ms Anna Rymszewicz School of Civil Engineering, UCD Dooge Centre for Water Resources Research University College Dublin Belfield Dublin 4 Ireland Email: [email protected]

Associate Professor Mary Kelly-Quinn School of Biology and Environmental Science, and UCD Earth Institute University College Dublin Belfield Dublin 4 Ireland Tel.: +353 1 716 2337 Email: [email protected]

Dr Liz Conroy School of Biology and Environmental Science, and UCD Earth Institute University College Dublin Belfield Dublin 4 Ireland Email: [email protected]

iii

Contents

Acknowledgementsii Disclaimerii Project Partners

iii

List of Figures

vii

List of Tables

x

Executive Summary

xiii

1 Introduction

1

2

4

3

4

Fine Sediment Sources, Delivery and Budgets  2.1

Sediment Sources and Delivery

4

2.2

River Bank Erosion as a Suspended Sediment Source

9

2.3

Deposition and Mobilisation

14

2.4

Construction Activities

14

2.5 External Discharges, Urban Drainage, Wastewater Treatment Plants and Farmyard Drains

15

Physical and Chemical Impacts of Fine River Sediments in Fluvial Systems

17

3.1

Importance and Processes

17

3.2

In-stream Processes

17

3.3

Sediment-associated Pollutants

21

3.4

Impacts on River Morphology

23

3.5

Impacts on Riverine Structures

24

3.6

Downstream Effects of Siltation in Rivers, Lakes, Reservoirs and Harbours

24

Ecological Impacts of Fine River Sediments in Fluvial Systems

27

4.1 Introduction

27

4.2

5

Periphyton and Macrophytes

27

4.3 Macroinvertebrates

31

4.4 Fish

33

Measuring and Monitoring Suspended Sediment Concentrations and Loads

35

5.1 Introduction

35

5.2

36

Manual Sampling

v

SILTFLUX Literature Review

6

7

8

9

10

5.3

Acoustic Doppler Current Profiler Method

38

5.4

Automatic Sampling for Suspended Sediment Concentration Time Series

38

5.5

Turbidimetric Instrumentation

40

5.6

Optical Backscatter Sensor Instrumentation

43

5.7

Laser In Situ Scattering and Transmissometry

44

5.8

Remote Sensing of Suspended Sediment Concentration

45

5.9

Estimation of Suspended Sediment Loads

45

Suspended Sediment Concentrations, Fluxes and Yields

49

6.1 Introduction

49

6.2 Ireland

49

6.3

51

Britain and Northern Europe

Storm-Event and Seasonal Suspended Sediment Dynamics

56

7.1

56

Storm-Event Suspended Sediment Dynamics

7.2 Hysteresis

56

7.3

Seasonal Changes in Suspended Sediment Fluxes

58

7.4

Longer Term Changes

60

Effects of Land Use and Climate Change on Sediment Fluxes

61

8.1 Introduction

61

8.2

Climate Change with Particular Reference to Ireland

61

8.3

Land Use

62

Management Implications

66

9.1

Reducing Sediment Load

66

9.2

Monitoring the Effectiveness of Measures

68

9.3

The Use of Modelling for the Design and Evaluation of Measures

68

Standards and Targets

69

References71 Abbreviations93

vi

List of Figures

Figure 1.1.

Suspended sediment in the River Alne, Warwickshire

2

Figure 1.2.

Suspended sediment of a very different colour (and probably source) in the River Alne, Warwickshire, at the same bridge site as shown in Figure 1.1, but looking upstream

2

Figure 2.1.

Example of an advanced classification system for potential hillslope and river channel suspended sediment sources

4

Figure 2.2.

Sediment connectivity in the fluvial system

5

Figure 2.3.

Natural and anthropogenic catchment and river processes that affect sediment dynamics

6

Figure 2.4.

Eroding arable fields – an example of a sediment source in Shropshire, UK

6

Figure 2.5.

The conceptual basis of the fingerprinting technique used to establish suspended sediment sources in the PSYCHIC study

7

Figure 2.6.

The conceptual framework that underpins the numerical INCA-Sed of Jarritt and Lawrence (2006)

8

Figure 2.7.

Sediment budget examples from catchments in central England

9

Figure 2.8.

River bank erosion on the River Allow, Ireland, is a sediment source

Figure 2.9.

Eroding river banks around a sedimentation zone on the River South Tyne, Northumberland11

11

Figure 2.10. World river bank erosion rates with respect to drainage basin area

12

Figure 2.11. River bank erosion as a sediment source in a reach-scale budget

12

Figure 2.12. River bank erosion events detected automatically with the PEEP system on the River Severn

13

Figure 3.1.

Turbid waters at high flow in the River Alne, near Little Alne, Warwickshire, UK

Figure 3.2.

Turbid conditions in the urban Bournbrook stream, River Tame catchment, Birmingham17

Figure 3.3.

Sediment infiltration mechanisms

18

Figure 3.4.

The hyporheic zone

19

Figure 3.5.

The relationship between stream power and sediment size in UK stream types: upland (Type I), small chalk (Type 2) and sandstone/limestone (Type 3)

19

Figure 3.6.

Downstream change in the hydraulic properties of the River Dart, southwest England

20

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17

SILTFLUX Literature Review

Figure 3.7.

Decline in oxygen supply rate with the accumulation of fine sediment within artificial redds

21

Figure 3.8.

Sediment pollution event in the nearshore zone derived from erosion of a coastal catchment during an intense Mediterranean rainstorm, east-central Spain, 24 August 1997

23

The continuum of channel planform variants of alluvial river morphology along an energy gradient is closely related to predominant sediment load and channel stability

24

Figure 3.9.

Figure 4.1.

Negative impacts of anthropogenically enhanced sediment on lotic aquatic systems27

Figure 4.2.

Schematic showing the mechanisms by which macroinvertebrates are affected (directly and indirectly) by suspended, deposited and saltating sediment particles

32

Figure 5.1.

Schematic of SSC cross-sectional variations

36

Figure 5.2.

Schematic cross-sectional variation in flow velocity, SSC and sediment flux

36

Figure 5.3.

Vertical distribution of concentration of various particle sizes in a stream section

37

Figure 5.4.

Time-integrating suspended sediment sampler for collecting large amounts of suspended sediment for composition analysis

39

Figure 5.5.

Infiltration basket for capturing fine sediment in gravel river beds

39

Figure 5.6.

Declining water clarity in Lake Tahoe, measured using the Secchi disk

40

Figure 5.7.

Turbidity versus SSC: calibration for the urbanised area at James Bridge, River Tame, Birmingham, UK

42

Figure 5.8.

Turbidity versus SSC: calibration for the large Skaftá river, south Iceland

43

Figure 5.9.

Dependence of light absorbance on sediment particle size

44

Figure 5.10. Effect of particle size on OBS response

44

Figure 5.11. Examples of SSC–Q relationships for two British rivers

46

Figure 5.12. Relationships between suspended sediment and area-weighted Q for several named British rivers

46

Figure 6.1.

Catchments that have been studied in recent sediment-related Irish studies

50

Figure 6.2.

Observed sediment yield (bedload and suspended load) data as a function of catchment area for UK rivers

52

Figure 6.3.

Downstream changes in optical water quality in six rivers in Wisconsin (USA) and New Zealand

55

Figure 7.1.

SSC response dynamics: SSC leading the flow and the classic positive hysteresis and first-flush model of sediment dynamics

57

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D. Lawler et al. (2010-W-LS-4)

Figure 7.2.

Classic suspended sediment dynamics in response to storm-event discharge changes on the River Dart, south-west England

58

Figure 7.3.

The typical suspended sediment dynamic response in the urbanised River Tame catchment (Birmingham, UK) is negative, anticlockwise hysteresis, in which peak SSCs occur just after the flow maximum

59

Clockwise hysteresis and anticlockwise hysteresis (the most common loops in the Q–turbidity relationship) for the River Tame, Birmingham

59

Figure 7.4.

ix

List of Tables

Table 1.1.

Initially proposed thresholds of SSCs for different effects on fish

1

Table 1.2.

Characteristics of fine sediments from selected UK rivers

3

Table 2.1.

Typical catchment sediment sources, and the likely variation in a downstream direction

5

Table 2.2.

Sediment sources for several south-west England catchments (delivery to watercourses in kg/ha per year), with information on source types for each catchment10

Table 2.3.

Summary of studies that have documented an increase in SSs downstream of river crossing construction sites

15

Table 2.4.

Major point sources of sediment

15

Table 3.1.

Selected examples of sediment-associated contaminants, their sources and their effects on fluvial systems

22

Table 3.2.

Bedform classification system

25

Table 3.3.

Alluvial depositional environments in which fine sediments may accumulate

25

Table 4.1.

The ecological impact and sources of suspended and deposited sediment in rivers

28

Table 4.2.

The effects of varying the concentrations of and the duration of exposure to suspended sediment on periphyton and macrophytes

29

Table 4.3.

The effects of varying the concentrations of and the duration of exposure to suspended sediment on macroinvertebrates

30

Table 4.4.

The effects of varying the concentrations of and the duration of exposure to sediment on fish

31

Table 6.1.

Summary of Irish sediment yields reported in the scientific literature

50

Table 6.2.

The Walling catchment typology: links to suspended sediment yield

53

Table 6.3.

The WFD catchment typology: links to suspended sediment yield

53

Table 6.4.

The new “Natural England” typology: links to catchment suspended sediment yield

54

Table 8.1.

Some effects of land cover changes on catchment characteristics

63

Table 8.2.

Total net rainfall, runoff and soil loss resulting from 30 storms between 28 May 1980 and 27 February 1981 in Nacogdoches, Texas

64

Table 9.1.

Reducing mobilisation of sediment from agricultural activities

67

Table 9.2.

Reducing delivery of mobilised sediment to watercourse

67

x

D. Lawler et al. (2010-W-LS-4)

Table 9.3.

Some estimated sediment reduction efficiencies

67

Table 9.4.

Reducing sediment export from urban areas to watercourses

68

Table 9.5.

Estimates of reduction efficiencies of best management practices for urban sediment

68

Table 10.1.

Proposed target and critical suspended sediment yields for various catchment types in England and Wales

69

Table 10.2.

Examples of standards/regulations for various countries

70

xi

Executive Summary

Sediment is a natural and dynamic component of river catchment systems, in which it is transported as bedload and/or suspended load, depending on the relationship between flow conditions, sediment supply and the structure, density, size and shape of materials. Although sediment does not feature explicitly within the Water Framework Directive (WFD), the ecological focus of the legislation with regard to surface waters means that the role of sediment as an essential component of the sustainable management of aquatic systems is recognised. Suspended sediment particles are typically < 63 μm in diameter, but can be much coarser (up to 2 mm in diameter in extreme events). The delivery of sediment to rivers is dependent on a range of factors including catchment soil type, vegetation cover, land use, hillslope hydrological processes, flow pathways, topographic setting and the presence/absence of buffer zones, such as floodplains, which can decouple hillslopes from river channels. In healthy fluvial systems, sediment provides the basis for diverse aquatic ecosystems through nutrient cycling and replenishment, as well as by forming the contributing materials from which aquatic habitats are constructed in river beds and especially in banks. Too much silt, however, can lead to the obstruction of channels, the smothering of habitats, ingress into the bed and the reduction of light penetration in the water column and at the bed, potentially leading to deoxygenation and environmental deterioration. The presence of elevated fine sediment may also degrade biological habitats through its biological oxygen demand (BOD) and the contaminants adsorbed to the sediments, and loss of habitat heterogeneity. Consequently, macroinvertebrate diversity and abundance is susceptible to change either directly, through effects on survivorship, growth, feeding, etc., or indirectly, through the alterations in habitat structure. Fish are also adversely affected by excessive sediment levels. Increased turbidity can cause gill irritation or reduce feeding activity by impairing visual range. Chemically active silt fractions (< 63 μm in diameter) can also act as important carriers of potentially hazardous nutrients and contaminants, dioxins and heavy metals. Sand deposition from the water column can

“seal” the surface of valuable gravel habitats, while silt particles (< 63 μm in diameter) can infiltrate the gravel matrix; these effects can significantly reduce the permeability and porosity of spawning gravels, suppress oxygen supply rates, hinder the removal of toxins from redds and reduce egg hatching and larval survival rates. Sediment infiltration can also negatively affect microbial processes in the hyporheic zones with consequences for biodiversity and ecosystem functioning, which may impact on groundwater quality. Therefore, the effective management of aquatic systems is critically linked to understanding sediment transport and storage pathways. While sediment is transported into river bodies through the natural processes of erosion and deposition, anthropogenic activities can also generate high sediment loadings. For example, human activities can lead to the deposition of soil as a result of the erosion of banks, due to trampling by livestock (known as “poaching”), the removal of riparian vegetation, the ploughing of land, deforestation/tree harvesting and land drainage schemes. The main route by which sediment is transported to water bodies is through drainage networks, but non-channelised surface wash flow can also be important, especially in arable contexts. Interflow and groundwater pathways can also be important in some settings. This review summarises the key issues that affect the role of fine sediment in fluvial systems, with a focus on northern Europe, the UK and Ireland, which will be of most relevance to the SILTFLUX project. The review includes definitions of fine sediment; descriptions of typical sediment sources and delivery mechanisms; overviews of the impacts of sediment (e.g. for organisms and downstream systems); descriptions of monitoring and analytical methodologies; details of fine sediment fluxes and yields for rivers in Ireland and comparable river catchments in the UK and elsewhere in Europe; a summary of the crucial dynamics of storm-event and seasonal suspended sediment transport; details of land use and climate change effects on sediment fluxes; and a discussion of the management implications of sediment.

xiii

1 Introduction

Sediment is an integral and dynamic component of healthy fluvial systems (Yarnell et al., 2006) and it plays a significant role in the geomorphological, hydrological and ecological functioning of a river (Kemp et al., 2011). However, its roles as a direct pollutant and as a vector of contaminant transport are now being increasingly recognised internationally (Ballantine et al., 2008; Collins et al., 2011). Indeed, excessive sedimentation is the number-one cause of water quality violations in the USA (Downing 2005). Metal and insecticide contaminants are also leading causes of violations, and most of these contaminants are bound to suspended sediment (Downing, 2008a). The management of sediment input to rivers is now a priority in many countries. This must recognise that the impact of sediment can be both due to suspended sediment in the water column and deposited sediment on the bed of river channels, lakes and estuaries. Many authorities have suggested maximum suspended sediment concentrations (SSCs), and Table 1.1 summarises some of these initial thresholds for fish. The Irish Environmental Protection Agency (EPA) has identified deteriorating water quality as the major environmental challenge and, in general, river pollution has increased since the late 1970s (Toner et al., 2000). Similarly, the England Catchment Sensitive Farming Delivery Initiative recognised “40 priority catchments in April 2006 where stakeholders require assistance to improve the protection of aquatic habitats” (Collins et al., 2010a). Catchment Sensitive Farming officers now assess contamination impacts, including those from sediment, and support stakeholders with the implementation of good practice on farms.

The aim of this literature review is to summarise the key issues affecting the role of fine sediment in fluvial systems, with a focus on northern Europe, the UK and Ireland, which are of most relevance to the SILTFLUX project. While the review primarily focuses on suspended sediments, deposited sediments are also considered. The review includes definitions of fine sediment; descriptions of typical sediment sources and delivery mechanisms; overviews of the impacts of sediment (e.g. for organisms and downstream systems); descriptions of monitoring and analytical methodologies; details of sediment fluxes and yields for rivers in Ireland and comparable catchments in the UK and elsewhere in Europe; a summary of the crucial dynamics of storm-event and seasonal sediment transport; details of land use and climate change effects on sediment fluxes; and a discussion of the management implications of sediment. Most of the data are contained in supporting figures and tables. There are already numerous and widely available international and European or UK-based literature reviews on suspended sediment transport, its impacts and associated pollutants, which have been published in the last 8 years and to which the reader is referred for details of specific issues (Collins and Walling, 2004; Lawler, 2005a; Owens, 2005; Owens et al., 2005, 2006; Walling et al., 2008; Gray and Gartner, 2009; Lawler et al., 2009; Taylor and Owens, 2009; Collins et al., 2011; Kemp et al., 2011; Owens and Xu, 2011; Vanmaercke et al., 2011). Therefore, reference is made to these sources as appropriate, and this review concentrates on the key issues relevant to the EPAfunded SILTFLUX research project.

Table 1.1. Initially proposed thresholds of SSCs for different effects on fish (Collins et al., 2012) Thresholds (mg/L) for least, probable and definite effects

Source

Least effects: high protection, best conditions

Probable effects: moderate protection, moderate conditions

Definite effects: low protection, poor conditions

< 25

25–80

> 80

EIFAC (1964)

< 25

26–80

> 80

Alabaster and Lloyd (1982)

< 30

30–85

> 83

Wilber (1983)

0

1–100

> 100

DFO (1983)

1

SILTFLUX Literature Review

In the context of this review, fine sediment (or suspended sediment) is defined operationally as sediment particles that are suspendable in the river water column (Figure 1.1 and Figure 1.2) and can ingress into gravel river beds; these particles typically range from 1 to 250 μm in diameter. For example, Walling et al. (2000) found that the particle diameter for > 95% of the suspended sediment load (SSL) sampled in most of the rivers in the Humber and Tweed basins in north-east England is < 63 μm (i.e. most are silt- and clay-sized particles). The median particle diameter, D50, ranged from 4.1 to 13.5 μm. Clay-sized particles (< 2 μm) accounted for 15–25% of SSL. A compilation of typical particle diameter distributions and organic content for bed and suspended sediment for several rivers in England and Wales is given in Table 1.2 (Buss, 2009).

Furthermore, Droppo (2003) argued that a broadening of suspended sediment definitions, from those that reflect purely physical characteristics to those that include biological and chemical information, is now required. For example, suspended sediment can often flocculate in aquatic systems if chemical and biological conditions are favourable, and these flocs will have different hydrodynamic properties, entrainment thresholds and settling characteristics (see, for example, Williams et al., 2008). The SSCs are normally measured in g/L or mg/L, while sediment fluxes are defined in, for example, kg per second (kg/s) or tonnes (t) per year. Longer term suspended sediment yields are usually normalised for catchment area to facilitate inter-basin comparisons and are presented as t/km2 per year. Field soil erosion rates are often given on the smaller scale of kg per hectare (kg/ha) per year (or t/ha per year).

Figure 1.1. Suspended sediment in the River Alne, Warwickshire. Flow from right to left (photo: Damian Lawler).

Figure 1.2. Suspended sediment of a very different colour (and probably source) in the River Alne, Warwickshire, at the same bridge site as shown in Figure 1.1, but looking upstream. Flow towards camera (photo: Damian Lawler).

2

3

< 250 μm surficial fine sediment

Little Stour, Kent, England

SD, standard deviation.

Fine bed sediment

Bed sediment

Suspended

Upper Piddle, Dorset, England

River Test, Hampshire, England

13.8 (SD 4.35, n = 51)

12.2

25–40 during summer and autumn low flows; 15–25 during winter and spring high flows

5–60

River Frome, Dorset, England

5.3 of < 2 mm 3.4 of < 2 mm

Suspended

Other

Spatially and temporarily consistent (D50 = 58.75 μm; SD = 6.25 μm)

Summer low flows (June– September): suspended sediment (< 0.25 mm) accounted for 70–90%; autumn floods (October): coarser sediment (0.25–4 mm) accounted for more

12.2% < 2 mm

28.9% < 2 mm

15.7% < 2 mm

1.7

0.6

0.6

Clay (< 0.0039 mm)

7.5 of < 2 mm

7.4

4.9

3.5

Silt (0.0004– 0.062 mm)

10% < 2 mm

45

85

23

Sand (0.063– 2 mm)

Particle size distribution (%)

19.7 of < 2 mm

Organic content (%)

River Blackwater, Hampshire, England

River Ithon, Powys, Wales

River Aran, Powys, Wales

River Test, Hampshire, England

Lowland limestone and sandstone streams

Accumulated sediment from artificial redd

Upper 30 cm of channel bed

Upland streams (impermeable strata)

Small chalk streams with low rainfall

Sediment type

River

Table 1.2. Characteristics of fine sediments from selected UK rivers (source: Buss, 2009)

Wood and Armitage (1999)

Walling and Amos (1999)

Acornley and Sear (1999)

Farr and Clarke (1984)

Greig et al. (2005)

Milan et al. (2000)

Reference

2

Fine Sediment Sources, Delivery and Budgets

2.1

Sediment Sources and Delivery

In addition, “information on sediment delivery to watercourses is urgently required to test and evaluate existing diffuse pollution models” (Collins et al., 2010a). Section 2.1.2 briefly summarises typical approaches adopted to identify or, rather, infer (Collins and Walling, 2004) catchment sediment sources, sediment routing and the processes that transport the sediment to the channel, that is, source-to-river connectivity.

2.1.1 Context Identifying sediment sources is extremely important. Collins and Walling (2004) have constructed a typology of likely sources (Figure 2.1), and argue that “an improved understanding of catchment suspended sediment sources represents a prerequisite (our italics) for assisting the design and implementation of targeted management strategies for controlling off-site sediment-associated environmental problems”. However, defining the provenance and types of fine sediment sources is very difficult and for many catchments these are largely unknown. Even within a given basin, dominant source locations may vary over time with different meteorological, hydrological and antecedent events. Typical fine sediment sources and how their relative dominance may change in a downstream direction in a catchment are listed in Table 2.1 for humid temperate environments. Such sources, however, must be viewed in contexts of the catchment and connectivity, as illustrated in Figures 2.2–2.4.

2.1.2

Methods of identifying and locating suspended sediment sources

Several direct and indirect approaches for estimating sediment sources exist, as reviewed by Collins and Walling (2004), and a combination of methods can be most effective. Sediment “fingerprinting”, through chemically “matching” likely source sediment with transported suspended sediment, has proved especially useful: its conceptual basis is shown in Figure 2.5. Multi-parameter or mixed model approaches to fingerprinting, as adopted by Walling et al. (2008) and Collins et al. (2010a) have become

Hillslope sources of suspended sediment Primary sources sheetwash rilling gullying trampling treethrow mass movement

Far

Near

redeposition

Secondary sources hillslope/ collovial deposits

remobilisation

remobilisation River channel sources of suspended sediment

Far

Links from hillslope to stream roads, paths, tracks, wheel ruts, drains

Near

Primary sources stream banks stream bed

redeposition

Secondary sources alluvial deposits

remobilisation

Instream Suspended Sediment Load Figure 2.1. Example of an advanced classification system for potential hillslope and river channel suspended sediment sources (after Collins and Walling, 2004).

4

D. Lawler et al. (2010-W-LS-4)

Table 2.1. Typical catchment sediment sources, and the likely variation in a downstream direction (adapted from Sear et al., 2003) Upper course

Middle course

Lower course

Rock fall

Valley side slope

Overland flow

Scree slope

Terrace slope

Tributaries

Debris flow

Soil creep

Cultivated farmland

Landslide

Floodplain erosion

Wind-blown soils

Freeze–thaw

Tributary stream

Construction sites

Sheet flow

Cultivated farmland

Urban runoff

Rills and gullies

Field drains and ditches

Gravel workings

Overgrazed, burnt or rabbit-infested areas

Urban runoff

Marine sediments (estuaries)

Ditches (forest and road)

Ditches (forest and road)

Quarries

Mining and gravel extraction

Figure 2.2. Sediment connectivity in the fluvial system (Crown copyright – Sear et al., 2003).

5

SILTFLUX Literature Review

Figure 2.3. Natural and anthropogenic catchment and river processes that affect sediment dynamics (Crown copyright – Sear et al., 2003).

Figure 2.4. Eroding arable fields – an example of a sediment source in Shropshire, UK (photo: Damian Lawler).

6

D. Lawler et al. (2010-W-LS-4)

Storm rainfall

Surface soils under woodland Surface soils under pasture

Mobilisation of sediment from individual sources

Surface soils under cultivation

Channel banks and subsurface sources of sediment

Mixing of mobilised sediment during sediment delivery

Suspended sediment load at catchment outlet

Comparison of catchment source material fingerprint properties of suspended sediment samples and individual catchment source materials.

Sediment source apportionment

Figure 2.5. The conceptual basis of the fingerprinting technique used to establish suspended sediment sources in the PSYCHIC study (adapted from Walling et al., 2008a). popular as methods of identifying catchment sediment sources.

because significant deposition of sediment usually occurs (1) before it reaches the drainage network and (2) within the river channel upstream of the sediment flux monitoring station. The percentage of sediment leaving a catchment, relative to that eroded from the catchment, is called the sediment delivery ratio (SDR), defined as:

There are a number of models for predicting sediment sources and sediment delivery from catchments, and linking them to in-stream sediment transport. A recent promising numerical model is the dynamic, process-based INCA-Sed (the Integrated Catchment Model for Sediments) (Jarritt and Lawrence, 2006). Its conceptual basis is shown in Figure 2.6. INCA-Sed works at a daily time step, and was tested by Lazar et al. (2010) in the River Lugg, a tributary of the River Wye in Wales. It was found that “diffuse soil loss” was the most important sediment generation process in the Lugg, although, in this case, SSCs were relatively low and unlikely to cause significant ecological impacts. It was also tested in four catchments in Finland by Rankinen et al. (2010) and was found to correctly reproduce the observed spatial and temporal sediment dynamics.

2.1.3

SDR = (100 SSY/E) × 100

(Equation 2.1)

In Equation 2.1, SSY is the fluvial suspended sediment yield per unit catchment area at some downstream river gauging point in the catchment (t/km2 per year) and E is the spatially averaged catchment erosion rate (t/km2 per year). In a small catchment SDRs can be as low as 15%, and these can be even lower in larger catchments, in which numerous sediment storage opportunities exist, at between 5 and 10% (Walling, 1983). Occasionally, a “hillslope SDR” is defined, which is the percentage of sediment eroded from slopes that reaches the river.

Sediment budgets and sediment delivery ratios

Quantifying the relative importance of different sediment sources allows a sediment budget to be constructed. A good example is provided by Walling et al. (2002) for the small lowland agricultural catchments of the Rosemaund and Lower Smisby in central

Measures of soil erosion rate cannot be used to infer rates of sediment delivery to stream channels

7

SILTFLUX Literature Review

splash erosion

infiltration rate

precipitation infiltration excess

detached sediment

surface runoff

point source inputs

flow erosion

sediment input to reach

saturation excess upstream import

soil water

water inflow to reach

suspended sediment

percolation

settling/ entrainment

bed sediment

downstream export

bank erosion

ground water

in- st ream phase Land p hase Figure 2.6. The conceptual framework that underpins the numerical INCA-Sed of Jarritt and Lawrence (2006) [after Lazar et al. (2010)]. England (Figure 2.7). Even in these tiny catchments (< 3.6 km2), the SDRs were only 17% and 20%, respectively, suggesting that the vast majority (≥ 80%) of the fine sediment produced in the catchment slope headwater areas was stored at locations up-catchment of the river monitoring sites, such as in fields and the channel (Figure 2.7), although field drains carried little sediment.

a shift towards more autumn-sown cereals which exposes bare tilled soils to the risk of erosion by winter rainfall and the expansion of maize production” (Collins et al., 2010a) (e.g. Figure 2.9). In the Midlands in England, Chapman et al. (2005) identified sub-surface drainage (macro­pore flow through cracked soils) as a key mechanism for transporting fine sediment from fields to streams. They also identified drains as a major mechanism for transporting fine sediment: in one of their catchments, drains contributed over half the annual sediment load.

Table 2.2 shows how multiple catchments can be separately assessed for their contribution to total sediment delivery, with additional evidence added for different source types (Collins et al., 2010a).

2.1.5 2.1.4

Sources and sediment pathways

Connectivity of catchments and channels

There must be good surface or sub-surface hydraulic connections that link eroding terrain parcels to the drainage network (i.e. strong connectivity) if abundant sediment is to reach the river channels (e.g. Figure 2.2). Indeed, Lane et al. (2007) showed, for the first time anywhere, that catchment topography and “hydrological connection discriminates both presence/ absence and abundance of juvenile brown trout populations”. This finding, based on work in the Eden catchment, Cumbria, could have wide ramifications elsewhere. For example, “if topographic control

Taylor and Owens (2009) compiled data that suggested that diffuse sediment from agricultural sources is responsible for 75.7% of all sediment supplied to rivers across England and Wales. Eroding channel banks (for an Irish example, see Figure 2.8) were responsible for a further 15.5% of sediment, with the remainder coming from urban sources (5.8%) and point sources (3.0%) (e.g. from wastewater treatment works). In the UK, arable fields can be key contributors of fine sediment: “sediment loss from cultivated fields has been accentuated by a number of factors including

8

D. Lawler et al. (2010-W-LS-4)

Gross surface erosion 466.6 t/km2

Rosemaund In-field 129.5 t/km2 Field drains

Subsurface erosion 4.5 t/km2

Field to channel 267.6 t/km2

Channel banks 9.4 t/km2

In channel storage 1.5 t/km2 Output 81.9 t/km2

Gross surface erosion 400.5 t/km2

Lower Smisby

In-field 112.4 t/km2 Field Drains

Subsurface erosion 2.3 t/km2

Field to channel 214.7 t/km2

Channel banks 5.5 t/km2

In channel storage 0.9 t/km2 Output 80.3 t/km2

Figure 2.7. Sediment budget examples from catchments in central England (Walling et al., 2002). mediates the watershed to stream linkage, land use impacts can only be appreciated with respect to their position in the landscape. This is of practical importance as locations of high connectivity should be a primary objective in targeting watershed restoration measures to where they will deliver most instream benefits” (Dugdale and Lane, 2006).

2.2

evidence from Church and Slaymaker, 1989; Collins et al., 1997; Lawler et al., 1999; Walling and Collins, 2005; Walling et al., 2008). World river bank erosion rates have been tabulated by Lawler (1993) and plotted against drainage basin area, as shown in Figure 2.10 (Lawler et al., 1997). Rates vary widely for a given catchment area, from 0.04 m/year for small catchments (< 10 km2) to 1 km/year for very large basins, such as the lower Mississippi (106 km2) (Figure 2.10).

River Bank Erosion as a Suspended Sediment Source

Bank sediment contributions can be modelled within a sediment budget framework [e.g. using the PSYCHIC model (Collins et al., 2007, 2009)]; however, in the field, this is challenging, even within a short river reach, as shown in Lawler et al. (1997) (Figure 2.11). This partly reflects the vast range of bank erosion processes possible (see below). It also reflects an

River bank erosion events can introduce significant quantities of fine sediment into stream channels (e.g. Figure 2.8 and Figure 2.9). Indeed, Collins et al. (2010a) argue that “there is increasing evidence for the role of eroding channel banks as an important sediment source in fluvial systems” (based on

9

10

229–236

33–40

Upper Parrett

Yeo

114–140

800–829

635–660

457–509

207–252

101–113

85–116

135–162

63–88

369–396

30–42

428–570

723–883

352–396

298–406

472–567

220–308

1292–1386

104–146

1496–1995

54–62

4–8

22–30

22–30

18–26

20–24

20–24

0–2

200

189–217

14–28

77–105

77–105

63–91

70–84

70–84

0–7

700

Damaged road vergesb

56–60

34–38

42–46

46–50

58–62

2–6

82–90

16–24

200

196–210

119–133

147–161

161–175

203–217

7–21

287–315

56–84

700

Channel banks/sub-surface sourcesb

Calculated using the total sub-catchment areas provided by the UK Environment Agency using the Flood Estimation Handbook.

Calculated using the area of the sub-catchment under the land use in question based on the ADAS land use database.

b

a

181–189

Tone

554–607

158–173

131–145

Isle

Parrett

511–567

327–376

146–162

94–107

Cary

Halse Water

591–647

169–185

700

200

200

700

Cultivated topsoilsa

Pasture topsoilsa

Sediment yield (kg/ha per year)

Brue

Subcatchment

2–6

2–6

0–4

2–6

4–8

0–4

2–6

0–2

200

7–21

7–21

0–14

7–21

14–28

0–14

7–21

0–7

700

Sewage treatment worksb

Table 2.2. Sediment sources for several south-west England catchments (delivery to watercourses in kg/ha per year), with information on source types for each catchment (adapted from Collins et al., 2010a)

D. Lawler et al. (2010-W-LS-4)

Figure 2.8. River bank erosion on the River Allow, Ireland, is a sediment source (photo: J.J. O’Sullivan).

Figure 2.9. Eroding river banks around a sedimentation zone on the River South Tyne, Northumberland (photo courtesy of the Northern Echo).

11

SILTFLUX Literature Review

Ilston sites

Published rates

Devon Streams

500

100

Bank Erosion Rate - Eb (m/year)

50

10 5

0.45

Eb = 0.0245 A

1 0.5

0.1 0.05

0.01 10

0

10

1

10

2

10

3

10

4

10

5

10

6

Drainage Basin Area - A (km2) Figure 2.10. World river bank erosion rates with respect to drainage basin area (adapted from Lawler et al., 1997).

qsin

Sediment supply from bank failure and lateral bank erosion

Input of sediment from upstream

qsbank Change in amount of sediment stored in reach s

qslat

Lateral exchange of sediment

Export of sediment to downstream

qsout s = qsin - qsout + qsbank + qslat

Figure 2.11. River bank erosion as a sediment source in a reach-scale budget (adapted from Thorne, 1991).

12

7

10

D. Lawler et al. (2010-W-LS-4)

inability, until recently, to monitor, in real time, the dynamics of the bank erosion processes that drive the specific bank erosion events that introduce fine sediment into streams. This makes it very difficult to assess, using modelling approaches, the linkage between river bank erosion and any resultant increase in fluvial SSCs.

shown for the Upper Severn in Wales in Figure 2.12 (Lawler et al., 1997). From this figure, it is apparent that the upper bank failed at midday on 24 August, while the lower bank has registered deposition, upon receiving the collapsed failure block at precisely the same moment. This not only helps to diagnose the bank erosion process responsible (i.e. a geotechnical failure in the case shown in Figure 2.12), but also indicates that at least some of the sediment released from the erosion of the upper banks may be stored at the bank toe, without immediately contributing significantly to fluvial SSLs. Blocks of detached fine sediment may lie at the bank foot for several months and leak sediment more slowly into the river. Furthermore, if the bank erosion process produces

These difficulties have been resolved, to some extent, by the development of the photo-electronic erosion pin (PEEP) system (Lawler, 1991, 2005b, 2008), which allows the magnitude, frequency, timing and duration of bank erosion events, and thus bank sediment contributions, to be monitored automatically and continuously. Examples of bank erosion events are

120

Reference cell

Cell series

Stage

60

Upper Bank

100

40 Stage (cm)

Voltage (mV)

80 60 40

20

20

20

21

22

23 24 Date in August 1993

25

26

60

100

Lower Bank

80

Voltage (mV)

0

40 60

40

Stage (cm)

0

20

20 0

20

21

22

23 24 Date in August 1993

25

26

0

Figure 2.12. River bank erosion events detected automatically with the PEEP system on the River Severn. Note that the upper bank shows bank erosion, while, at the same time, the lower bank shows deposition; this suggests a geotechnical bank failure/block collapse process (adapted from Lawler et al., 1997).

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SILTFLUX Literature Review

fine sediment particles or aggregates, perhaps through freeze–thaw and desiccation processes followed by fluvial entrainment, then erosion processes can immediately and significantly elevate SSCs. Lawler et al. (1997) suggested catchment-scale differences in bank erosion process dominance, namely that freeze–thaw processes dominate in the upper reaches of a catchment, where banks are wetter and subject to more frost; fluid entrainment dominates in the middle reaches, where stream powers may peak (Barker et al., 2009); and in the lower reaches, where banks are large and fine grained, geotechnical failure is the most common cause of erosion. Therefore, it is possible that, in cases where bank erosion processes are significant contributors to the sediment budget, more immediate suspended sediment responses are likely in the middle and upper reaches of catchments.

2.3

1990). The original Shields diagram used a time and spatially averaged dimensionless shear stress and a particle Reynold’s number calculated at the bed. However, it does not apply to sediment particles of low specific gravity and small diameter. Mantz (1977) extended the Shields diagram for smaller particles, but both approaches require iteration. Yalin (1977) introduced a change of variables, defining a modified Shields function, which could be used without the need for iteration. The movement of relatively large-grained sediments tends to be associated with less frequent but more intense flows, and can be modelled using typical statistical distributions of extremes (Valyrakis et al., 2011). Entrainment is influenced by the velocity profile in the boundary layer (Le Roux, 2010). Scour may occur at specific structures, such as flow deflector vanes used to improve conditions for fish (RodrigueGervais et al., 2010), arched culverts (Crookston and Tullis, 2011) and weirs dal (Muller et al., 2011), which change the flow direction or increase its velocity.

Deposition and Mobilisation

Flowing water can detach, transport and deposit sediment, whether in overland flow or in channels. Critical shear stress is an important index which characterises this process (Storm et al., 1990). The shear stress required for detachment is typically much greater than required for transport: Foster (1982) reported that the critical shear stress required for the flow detachment of a certain soil was 2.9 N/m2, while the critical shear stress required for the transport of the same soil was <0.5 N/m2. Critical shear stress is related to soil shear strength in cohesive soils and is influenced by salinity (Kelly and Gularte, 1981). Lyle and Smerson (1965) reported that critical shear stress is related to plasticity index, percentage clay, mean particle size, dispersion ratio, vane shear strength, organic matter content, cation exchange capacity and the calcium–to–sodium ratio. Smerdon and Beasley (1961) developed regression equations relating critical shear stress to many of these soil properties. In channels, sediment can move, suspended in the turbulent water column, or can roll downstream on the bed of the channel. The latter is called “bed-load” and Yalin (1963) developed an equation to describe the transport rates. Sediment moving along the bed may form characteristic waves and these can influence the hydraulic resistance of the channel (Wang et al., 2011).

2.4

Construction Activities

Road construction activities in the vicinity of watercourses are potential sources of sediment input, which may originate from the associated earthworks, including blasting; the pumping of water from the construction site; exposed soil banks resulting from excavations or vegetation removal; soil storage areas; or the construction of road crossings (Vice et al., 1969; Barton, 1977; Beshta, 1978; Extence, 1978; Cline and Forest, 1983; Embler and Fletcher, 1983; Duck, 1985; Ellis et al., 1987a; Barrett et al., 1995; Maltby et al., 1995; Luce and Black, 1999; Wellman et al., 2000; Lane and Sheridan, 2002; Bruen et al., 2006; Cerdà, 2007; Purcell et al., 2012). Pipeline crossings are especially troublesome zones for fine sediment ingress into rivers and fluvial sediment impact assessment methodologies have recently been presented by Lawler and Wilkes (2015) for these types of impacts. The risk from the construction of road crossings is also high; the available data on road crossings and suspended solids (SSs) are summarised in Table 2.3. Culverting appears to pose a higher risk of sediment than the use of clear span bridges (Cocchiglia et al. 2012). The limited studies that relate these sediment inputs from road crossing to effects on aquatic

The most widely used method for determining critical shear stress is the Shields diagram (Storm et al.,

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D. Lawler et al. (2010-W-LS-4)

Table 2.3. Summary of studies that have documented an increase in SSs downstream of river crossing construction sites Crossing type

SS concentration (mg/L) before construction/ control

SS concentration (mg/L) during construction

Note

Reference

Culvert

< 5

1390

Maximum values

Barton (1977)

Pipeline

7

7620

Maximum

Tsui and McCart (1981)

Bridge foundations

3.2

15.8

Mean

Cline et al. (1983)

Culvert

3–17

75–81

Range

Cline et al. (1983)

Culvert

< 30

60–130

Range

Embler and Fletcher (1983)

Unknown

35

179

Mean

Barrett et al. (1995)

Culvert

144

1237

Maximum values

Lane and Sheridan (2002)

Unknown

5

15

Maximum values

Chen et al. (2009)

Culvert

< 20

70

Maximum

Purcell et al. (2012)

communities include Extence (1978) and Cocchiglia et al. (2012). Post-construction, sediment and various other pollutants have been detected in motorway drainage (Maltby et al., 1995, e.g. Bruen et al., 2006a).

2.5

Much of the sediment can be contained in settling and storage tanks on site, particularly if tanks are designed to accommodate the “first flush” load (Walling, 2006). However the organic content and nutrients in wastewaters may have a greater effect than sediment on stream ecology (Horowitz, 1991).

External Discharges, Urban Drainage, Wastewater Treatment Plants and Farmyard Drains

Sediment concentrations in urban stormwater depend on whether the urban area is residential, commercial or mixed. Median event mean concentrations tend to be in the range of 100 to 300 mg/L, but values as high as 1240 mg/L have been observed in residential areas (Van Rompaey et al., 2001). Total suspended solids (TSSs) are typically reduced in stormwater by detention or settling ponds (Carter and Berg, 1983; Sarangi et al., 2004; Richards et al., 2008). The process of urbanisation initially produces a large increase in sediment load as land is cleared for construction (Wolman, 1967). This sediment load will decline as more impermeable surfaces are built. However, the reduction in the sediment load in the stream may result in channel erosion, and a deepening

Most of the sediment input to rivers from point sources is associated with local rainfall events, typically with the intense but relatively short-duration rainstorms. Such point sources include urban stormwater drainage, sewer or wastewater treatment plant (WWTP) overflows, road runoff and farmyard drainage (Table 2.4). Sediment delivery from WWTP discharges, particularly in areas with combined sewers, and urban stormwater drainage systems is complicated by the effects of deposition within the pipe networks and subsequent mobilisation during storm flows (Walling et al., 2003a).

Table 2.4. Major point sources of sediment Point source

Reference

Urban stormwater drainage

Carter and Berg (1983), Sarangi et al. (2004), Richards et al. (2008)

WWTP discharges

Walling et al. (2003a)

Road runoff

Walling (1983), Croke et al. (2006), Desta et al. (2007)

Farmyard drainage

Van Oost et al. (2000), Krasa et al. (2010)

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SILTFLUX Literature Review

and widening of the urban channel, if the bed material permits (Leopold, 1973). Some examples from the USA are listed by O’Driscoll (2010).

increased over time (Collins et al., 2010b). A “first flush” effect of heavy metals is typical of runoff from busy highways (Walling, 1983), regardless of whether the road is unsealed (Croke et al., 2006) or sealed (Ellis et al., 1987b). Bruen et al. (2006) reviewed the typical constituents of highway runoff and their impacts, and Desta et al. (2007) described a study in Ireland, in which all TSS measurements in road runoff were considerably less than 1000 mg/L. In rural areas, the sediment loads in farmyard runoff can be reduced by deposition in either natural or constructed wetlands (Van Oost et al., 2000; Krasa et al., 2010).

Roads are sources of sediment in runoff. In urban areas, the runoff is mainly to the urban stormwater system, but in rural areas it discharges to local streams, perhaps after some treatment in, for example, a settling/attenuation pond or wetland (Schutes et al., 2001). In one UK study, damaged road verges contributed up to 20% of the sediment from catchments, and this contribution seems to have

16

3

Physical and Chemical Impacts of Fine River Sediments in Fluvial Systems

3.1

Importance and Processes

SSCs in the water column; the particle size and shape distribution; particulate behaviour and ingress into river bed gravels; sediment quality; sediment-associated contaminants; sediment fluxes; and storage characteristics/dynamics in the floodplain (Walling et al., 1999; Coulthard and Macklin, 2003; Walling et al., 2003b; Dennis et al., 2009), channel bed and hyporheic zone.

Despite the potential negative physical impacts of suspended sediment (the ecological impacts are discussed in Chapter 4), a little suspended sediment can limit water clarity and, therefore, the overproduction of algae. Fluvial suspended sediment data can also help to identify catchment processes, such as: 1. upstream sediment mobilisation and hence data that can provide alerts to potential erosion problems, including the liberation and delivery processes (e.g. soil erosion, rilling and gullying, and landslide and debris flow significance) that route fine sediment to downstream receptors;

3.2

In-stream Processes

Once in the channel, and especially once suspended in the water column in sufficient quantities, particulate matter can make the water turbid, at both high (Figure 3.1) and low flow, especially in urban settings (Figure 3.2). This reduces the level of light that can reach the bed and affects aquatic habitats and organisms, especially predation for fish (Walling and Fang, 2003), as discussed in Chapter 4. Methods for modelling the factors that control photosynthetically active radiation in rivers, including turbidity, have recently been developed using the Benthic Light Availability Model (Julian et al., 2008a,b).

2. catchment erosion and sediment yields (e.g. for reservoir design purposes); 3. downstream sedimentation (e.g. in lakes, nearshore zones and harbours); 4. sediment fluxes and any associated delivery of contaminants to receptors. However, the presence of fine sediment in catchment and fluvial systems can also have undesirable environmental impacts, because sediment can change the physical, chemical and biological properties of aquatic ecosystems. These impacts depend on the

Sediment can also have physical abrasive “sandblasting” effects on organisms (e.g. damage to fish gills; see section 4.4) and on riverine structures, such as bridges, intakes and turbines. In addition, sediment physically interacts with the river channel

Figure 3.2. Turbid conditions in the urban Bournbrook stream, River Tame catchment, Birmingham (photo: Damian Lawler).

Figure 3.1. Turbid waters at high flow in the River Alne, near Little Alne, Warwickshire, UK. Flow from right to left (photo: Damian Lawler).

17

SILTFLUX Literature Review

in a number of other ways (Lawler and Fairchild, 2010). “Sedimentation” (or “siltation”) refers to the development of a layer of fine sediment over the bed surface. “Sediment infiltration” (or “colmation”, as it is known in the environmental engineering literature) is the process by which fine sediment moves into the gravel bed structure itself (Figure 3.3). “Accumulation” is the summation of this infiltration process over time (Sear et al., 2008).

bed by fluid turbulence. All else being equal, coarser and heavier particles will drop out of suspension first, giving natural spatial and temporal size segregation in the resulting deposits. Particle shape is also a key factor, as the less spherical a particle, the slower it will settle.” If fine particles flocculate they can settle in an unpredictable manner. The hyporheic zone is the zone in which groundwater and surface water mix beneath and around the channel bed (Alley et al., 2002) (Figure 3.4). This zone is especially vulnerable to sediment ingress problems (see, for example, Sear et al., 2008; Lawler et al., 2009). This partly relates to a difficulty in flushing out fine sediment once it has become ingressed into the bed gravels. Data from a sample of UK stream beds,

In the Hyporheic Handbook, Lawler et al. (2009) briefly summarise the processes involved in sediment infiltration: “in the water column, fine sediment movements are driven by two main processes: (i) gravity-driven infiltration that includes simple Stokestype settling; and (ii) advection of fine material into the

Figure 3.3. Sediment infiltration mechanisms (Sear et al., 2008).

18

D. Lawler et al. (2010-W-LS-4)

watertable

Stream

Groundwater flow

ble

erta

wat

Groundwater flow

Hyporheic Zone

Figure 3.4. The hyporheic zone.

Percentage sized under 1 mm

100

10

1 1

10

100

1000

10000

Total Stream power (Wm ) -2

Upland

Sandstone /Limestone

Small chalk

Figure 3.5. The relationship between stream power and sediment size in UK stream types: upland (Type I), small chalk (Type 2) and sandstone/limestone (Type 3) (after Milan et al., 2000). shown in Figure 3.5, suggest that there is a reasonably inverse relationship between specific stream power (a measure of the energy available per unit area of a river bed) and the percentage of sediment of < 1 mm in diameter (Milan et al., 2000). If this is the case in general, it should be possible to use the highresolution (60 m longitudinal spacing) catchment-scale quantifications of downstream changes in channel elevation, slope, bankfull discharge, and gross and specific stream power, produced for 32 UK rivers using the Combined Automated, Flood, Elevation and Stream Power (CAFES) system of Barker et al. (2009), as a basis to predict the fine sediment content of river

beds (e.g. Figure 3.6). The resulting data could then be employed to help identify, in combination with other indices, hot-spots of ecological sensitivity. Fine sediment in gravels can also consume oxygen and, therefore, significantly reduce organisms’ oxygen supply (Greig et al., 2007) (see section 4.3.1). Figure 3.7 shows that a simple doubling of sediment accumulation rate in artificial redds can reduce oxygen supply by an order of magnitude (Greig, 2004). Sedimentation also changes channel geometry, and if reductions in channel capacity are significant, then flow velocities (and shear stresses) will increase to

19

SILTFLUX Literature Review

Figure 3.6. Downstream change in the hydraulic properties of the River Dart, south-west England. Sites of erosion are likely to relate to peak stream power, which in many catchments occurs in mid-basin, in which the optimum combination of water surface slope and bankfull discharge [median annual discharge (QMED), i.e. 2-year return period flow] is found, as in the River Dart in this example. The full derivation of variables was performed using the CAFES methodology (Barker et al., 2009). The results shown here are from Barker et al. (2009). Sediment deposition and associated impacts are likely at stream confluences and in areas of low stream power, at which particle settling velocities are achieved (Buss, 2009).

20

D. Lawler et al. (2010-W-LS-4)

1000 100

Test Redd 2

10

Ithon Redd 1

2

Oxygen supply rate (mg O egg

-1

hour

-1

)

Test Redd 1

Ithon Redd 2 1 Blackwater Redd 1 0.1 0.01 0.001 0.0001

0

20

40

60

100

80 -2

Total sediment accumulation (kg m

)

Figure 3.7. Decline in oxygen supply rate with the accumulation of fine sediment within artificial redds. These data are based on Greig, 2004. balance the continuity equation. This can be beneficial in some cases (e.g. to flush out accumulated fine sediment), but it can also increase channel erosion and bedform instability, perhaps undesirably.

than 90% of pollutants are transported with sediments and, more recently, Turner et al. (2008) showed that between 71% [copper (Cu)] and 99% [arsenic (As), lead (Pb) and zinc (Zn)] of selected heavy metals were sediment bound in waters buffered by calcareous soils, under high wash sediment loads (up to 10,000 mg/L). Significant amounts of the nutrients nitrogen and phosphorus have also been widely reported for sediment-borne fluxes in British and Irish catchments (see, for example, Foster et al., 1996). However, in a review of the controls on nutrient fluxes in British rivers, Walling et al. (2001) emphasised that the “precise magnitude of the sediment-associated component will vary from river to river in response to local conditions including relief, geology and land use, the hydrometeorological conditions and the relative importance of point source inputs”. Similarly, at a reach scale, the morphological controls governing sediment dynamics strongly influence the transport and storage of radiothorium downstream of a mining release (Graf et al., 1991), while Droppo et al. (2011) have demonstrated that bacterially induced fine sediment flocculation can lead to an increase in the “downwards flux” of contaminated sediment, which promotes elevated pathogen distribution in bed sediments at a patch scale (see, for example, Cho et al., 2010).

Because of these negative impacts, and those discussed in Chapter 4, organisations around the world have established limits for SSCs, turbidity and/ or sediment fluxes. For example, the UK Technical Advisory Group (UKTAG, 2008) suggested that the guideline standard for annual mean SSs in freshwaters should be 25 mg/L; this is consistent with the EU Freshwater Fish Directive (Lazar et al., 2010). The EU Water Framework Directive (WFD) required EU Member States to achieve “good ecological status” for rivers by 2015, and this status was defined by Collins et al. (2007) in terms of the guideline annual average SSC of 25 mg/L of the EU Freshwater Fish Directive. Other suggested SSC thresholds are given in Table 1.1.

3.3

Sediment-associated Pollutants

The role of sediments in the transport and dispersal of river pollutants has received considerable attention since the 1970s. Gibbs (1973) and Martin and Meybeck (1979), for example, suggested that more

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There have been numerous reviews and case-study papers on the variety and source of sedimentassociated pollutants’ distribution and effects in aquatic environments; a number of these are summarised in Table 3.1. Notably, relatively few investigations have been conducted in Ireland, although Herr and Gray (1997), and a follow-up study by Gaynor and Gray (2004), did report elevated levels of sedimentassociated Cu, Pb and Zn as a result of acid mine drainage in the Avoca catchment, south-east Ireland, and Regan et al. (2012) have estimated phosphorus exports of 0.88 to 8.8 kg/ha per year from agricultural soils in Ireland, based on total phosphorus soil values (after McGrath et al., 2001) and soil erosion modelling (Van Oost et al., 2006).

Pollutants can be introduced as particulates that move as part of the saltating load, but are more often transported in association with the “chemically active”, fine (< 63-µm diameter) sediment fraction, in suspension (Horowitz, 1991), through a variety of processes (adsorption, absorption and precipitation) (see Figure 3.8). Sediment-associated pollutant concentrations are therefore strongly influenced by particle size, while total pollutant fluxes are primarily governed by pollutant mobilisation, coupled with sediment supply and hysteresis dynamics (Bradley and Cox, 1990); thus, very significantly, most sediment-associated pollutant transport occurs during storm events (Horowitz et al., 1999; Horowitz, 2006; Edwards and Withers, 2008). The disassociation of pollutants from the dissolved,

Table 3.1. Selected examples of sediment-associated contaminants, their sources and their effects on fluvial systems (adapted from NRC, 2003; Owens et al., 2005; Taylor and Owens, 2009). Brief details of research conducted in Ireland are given in Italics Pollutant

Sources and distribution

Environmental impacts

Selected examples

Metals and metalloids (Ag, Cd, Cu, Co, Cr, Hg, Ni, Pb, Sb, Sn, Tl, Zn, As)

Mining including catastrophic release, industry, acid rock/mine drainage, sewage treatment, urban runoff, pesticides

Toxicological effects

Herr and Gray (1997)/Gaynor and Gray (2004) (Ochre precipitates in the Avoca River)

Oxidisation leading to deoxygenation Potentially long residence times (102–103 years)

Miller (1997), Macklin et al. (2003), Walling et al. (2003b)

Diffuse source of labile contaminants and prone to physical reworking Nutrients (P, N)

Organic compounds including pesticides, herbicides, hydrocarbons, PCBs, PAHs, dioxins

Agricultural and urban runoff, wastewater and sewage treatment

Eutrophication

Agriculture, pest control, industry, sewage, landfill, urban runoff, forest fires, incineration

Toxic impacts, including carcinogenesis

Regan et al. (2012) (Review of P exports from tillage) Kiely (2007) (P exports from grasslands) Thomas (1990), Potter et al. (1994), Warren et al. (2003)

Endocrine disruption affecting sex hormones and reproductivity Environmental persistence and bioaccumulation

Xenobiotica and antibiotics

Sewage treatment works, industry, agriculture

Steroid hormones (e.g. androgens and oestrogens)

Sewage treatment works

Radionuclides (137Cs, 238Pu, 239 Pu, 240Pu, 235U, 238U, 230Th, 99 Tc)

Nuclear power industry, military, geology, food irradiation

Microbes, including pathogens (Escherichia coli, faecal coliforms)

Agricultural runoff, sewage, landfill

Endocrine disruption

Taylor and Harrison (1999)

Formation of flocs leads to bed concentration of pathogens

Jamieson et al. (2004), Cho et al. (2010), Droppo et al. (2011)

Ag, silver; Cd, cadmium; Co, cobalt; Cr, chromium; Cs, caesium; Hg, mercury; N, nitrogen; Ni, nickel; P, phosphorus; PAH, polycyclic aromatic hydrocarbon; PCB, polychlorinated biphenyl; Pu, plutonium; Sb, antimony; Sn, tin; Tc, technetium; Th, thorium; Tl, thallium; U, uranium.

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Figure 3.8. Sediment pollution event in the nearshore zone derived from erosion of a coastal catchment during an intense Mediterranean rainstorm, east-central Spain, 24 August 1997 (photo: Damian Lawler). aqueous phase generally reduces their bioavailability and environmental impact. However, fine sediments continue to represent a risk to the environment through their ingestion and, importantly, represent a major secondary source of pollution. The diffuse nature of the sediment-associated pollutants stored in the channel, and the channel marginal and floodplain sedimentary environments, makes them more difficult to manage (Owens et al., 2005). Moreover, some pollutants with long residence times (up to 1000 years) can be re-introduced to surface waters, through chemical exchange and channel re-working, long after their initial release into the fluvial environment (Miller, 1997; Macklin et al., 2003; Dennis et al., 2009). In many instances, keeping rivers “clean” of fine sediment should, therefore, enhance their ability to flush through pollutants, leading to significant environmental benefits in the medium term.

3.4

(Brown and Brussock, 1991; EA, 2009). Sediment transfer also promotes short- and long-term storage of fine-grained sediment in a variety of alluvial depositional settings (Table 3.3). Typically, channel form–sediment dynamics are dominated by bedload sediment characteristics, but fine sediment can be an important constituent of mixed bed channels, in which it is stored in the interstices of coarser channel substrate and deposited on bar tops during waning high flows (Church, 2006). Fine sediment can build up (sometimes detrimentally) in channel pools and in the lee of in-channel obstructions, such as woody debris and man-made structures. In the upper portion of alluvial channel banks, fine sediment plays an important morphological role, because of its hydraulic and cohesive properties, and provides a store for nutrients and, in some cases, contaminants. Fine sediment predominates in the alluvial sedimentary environments of low-energy systems. Morphologically, these river systems are typically laterally stable and can display anabranching, anastomosing (sensu Nanson and Knighton, 1996) planforms, characterised by multiple, hydraulically independent channels separated by stable, vegetated floodplain islands. These equilibrium channel planforms are now rare, because of human disturbance and modification, but rudimentary morphological elements in the form of channel islands can still be found across north-west Europe, including in Irish catchments such as those

Impacts on River Morphology

Form–sediment interactions can result in distinctive channel planforms (Figure 3.9) and bedforms (Table 3.2), which are largely governed by sediment calibre and supply rate, together with local flow conditions. Although sediment moves through a channel, bedform assemblages, such as riffle-pool sequences, are equilibrium forms that are morphologically stable and provide habitats for aquatic species

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Decreasing grain size and stream power -->

gravel bed braided

Increasing Energy --->

boulder bed

sand bed braided

wandering sand bed

wandering gravel bed

sand bed meandering

fine-grained meandering

low sinuosity fine-grained

anastomosing fine-grained

Figure 3.9. The continuum of channel planform variants of alluvial river morphology along an energy gradient is closely related to predominant sediment load and channel stability (after Brierley and Fryirs, 2005, p. 121).

3.6

of the Rivers Shannon (County Limerick) and Boyne (County Meath), and are highly developed in parts of the River Lee (County Cork).

3.5

Downstream Effects of Siltation in Rivers, Lakes, Reservoirs and Harbours

Siltation processes can have a significant impact on fluvial systems and water resource management in river catchments, and can also account for the low SDRs reported for many river basins (see section 2.3). Fine sediments are temporarily stored in-channel, in substrate interstices and bar tails (section 3.5). Artificial channels, such as drainage ditches (also known as field drains), both capture and supply fine sediment from agricultural land (Forster and Abrahim, 1985). An increase in the drain density will improve hydrological and sediment connectivity, but can lead to elevated particulate phosphorus transfer to aquatic ecosystems (Blann et al., 2009) and high sediment yields.

Impacts on Riverine Structures

Suspended sediment particulates can be sharp and angular, and abrade in-stream structures such as bridges, flood defences, hydroelectric power turbines (e.g. in Iceland and Norway) and dam surfaces, particularly spillway chutes (Huang, 2006). They can also clog water intakes, creating problems in water treatment works and increasing the costs of sediment removal. Floodplain and main channel sedimentation can also become problematic. Turbulent diffusion can contribute greatly to sediment transport (Shiono, 2011), and the abrasion of coarse sediment during transport may produce fine silt (Sklar et al., 2006).

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Table 3.2. Bedform classification system (modified from Church and Jones, 1982a, and Knighton, 1998) Bedform

Dimensions

Small-scale forms

(10 –10 m) –2

Shape

Behaviour and occurrence

2

(1) Sand-bed rivers Ripples

λ < 0.6 m; H < 0.04 m

Triangular profile, sharp crest perpendicular to flow direction

Generally restricted to sediment finer than 0.6 mm

Dunes

λ is 4- to 8-fold higher than flow depth; H is up to 1/3 of flow depth

Curved crests

Upstream slope may be rippled, coarse grains deposited at crest, flow separation occurs

Flat surface

Super critical flow, Froude No. > 1

H is dependent on flow depth/velocity

Sinusoidal profile

Antidunes move upstream, sediment moves downstream

Pebble clusters

0.1–1 m

Linear in direction of flow

Large obstacle particle with smaller stoss particle

Transverse ribs

1–10 m

Transverse to direction of flow

Repeated ridges of coarse particles, spacing proportional to largest particle in ridge crest

Riffle/pool sequence

1–10 m

Transverse to direction of flow

Alternative deep (pool) and shallow (riffles) spaced in relation to channel width

Step/pool systems

1–10 m

Transverse to direction of flow

Star-like sequence formed in steep channels, steps formed coarse material, spacing ≈2- to 3-times the channel width

Variable

Longitudinal bars, form in channel centre and elongated in direction of flow

Plane bed Antidunes (2) Gravel-bed rivers

Large-scale forms

(101–103 m)

Bars

Length comparable to channel width

Transverse bars, lobe shape with relatively steep downstream face Point bars, form in meandering channels because of secondary flow Diagonal bars, bank attached bars running obliquely across channel Mid-channel bars, common in braided channels λ, wavelength; H, height.

Table 3.3. Alluvial depositional environments in which fine sediments may accumulate (modified from Church and Jones, 1982b, and Hoey, 1992) Depositional site

Name

Characteristics

Within channel

Transient channel deposit

Temporarily stagnant bedload deposits including ripples, dunes, transverse ribs, pebble clusters and steps

Alluvial bars

Lag deposits of coarser sediment including riffles, mid-channel bars and sedimentation zone

Channel margin

Lateral deposits

Point bars that form on the inside of meander bends

Floodplain

Vertical accretion deposits

A comparatively flat alluvial depositional landform that forms as a result of deposition of fine-grained suspended load of overbank floodwaters; provides sediment storage space for drainage basin

Piedmont

Alluvial fans

A cone-shaped depositional feature that occurs because of a sudden reduction in sediment transport capabilities owing to an abrupt change from confined to unconfined conditions, or a sudden decrease in slope; sediment grain size decreases rapidly with distance from fan apex

River mouth

Deltas

Morphological feature formed when a river enters a sea or lake and deposits its load; characteristics of sediment supply determine morphology

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SILTFLUX Literature Review

Fine sediments naturally accumulate in natural catchment sinks, such as lakes, flood basins and estuaries, but can be affected by changes in sediment supply. For example, Lake Tahoe, on the California/ Nevada border, has traditionally attracted many tourists; however, this tourism is being threatened by an apparent increase in turbidity resulting from an increase in the sediment and nutrient supply (Langlois et al., 2005; Simon et al., 2008). In artificially constructed reservoirs and impounded rivers, the effects of siltation can also have a detrimental impact on storage capacity, particularly if trap efficiencies are underestimated; this impacts upon a variety of uses, such as flood control, domestic water supply,

irrigation, recreation and fish farming. Siltation can also exert pressure on dam walls, increasing the risk of dam failure; while these events are rare, they can have catastrophic downstream effects, because of the mobilisation of trapped fine sediment (e.g. Evans, 2000, 2007). Harbours and marinas often suffer from fine sediment accumulation, which leads to high maintenance costs for dredging and the disposal of spoil, particularly if it is contaminated (Owens, 2005). These effects can result from the downstream deposition of fine sediment (Mitchell, 2005; Pontee and Cooper, 2005), together with the re-suspension of fine sediment from tidal flats (Dobereiner and McManus, 1983).

26

4

Ecological Impacts of Fine River Sediments in Fluvial Systems

4.1 Introduction

4.2

Excessive sediment inputs can have detrimental effects on fluvial systems, negatively impacting biota in a variety of complex direct and indirect ways (Figure 4.1), from reducing primary productivity to altering faunal abundance, diversity and community structure. The key papers documenting such effects on aquatic biota are summarised in Table 4.1, whilst Table 4.2, Table 4.3 and Table 4.4 summarise the available studies that have considered the influence of sediment concentration and the duration of exposure.

Periphyton and Macrophytes

4.2.1 Periphyton Periphyton is, to some extent, more susceptible to effects resulting from elevated sediment levels than macrophytes. Suspended and saltating solids alter light penetration within the water column, and thereby reduce primary production (Van Nieuwenhuyse and LaPerriere, 1986; Lloyd et al., 1987) and lead to a decrease in periphyton biomass (Davies-Colley et al., 1992; Quinn et al., 1992; Francoeur and Biggs, 2006;

Figure 4.1. Negative impacts of anthropogenically enhanced sediment on lotic aquatic systems. Rectangles represent physicochemical effects, ovals represent direct and long-term biological and ecological responses (after Kemp et al., 2011). BOD, biological oxygen demand.

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Table 4.1. The ecological impact and sources of suspended and deposited sediment in rivers (modified from Wood and Armitage, 1997, and Jones et al., 2011) Impact

Sediment type

Source

Reference(s)

Absence of rooted vegetation

D

China clay extraction

Nuttall and Bielby (1973)a

Elimination of macrophytes

D

Channelisation

Brookes (1986)a

Reduced productivity, biomass and organic content

S and D

Placer gold mining

Van Nieuwenhuyse and LaPerriere (1986), Davies-Colley et al. (1992)b,c

Reduced species diversity and organic content

S and D

Road construction

Cline et al. (1982)b

Reduced organic content

D

Impoundment

Graham (1990)c

Reduced abundance

D

Drought – abstraction

Wright and Berrie (1987), Wood and Petts (1994)a

Reduced abundance and diversity

D

China clay extraction

Nuttall (1972), Nuttall and Bielby (1973)a

Reduced abundance and diversity

S and D

Desilting operation

Doeg and Koehn (1994)e

Reduced species diversity

D

Removal of riparian vegetation

Armstrong et al. (2005)e

Reduced species density

S and D

Placer gold mining

Quinn et al. (1992)c

Reduced species density

S and D

Induced in field experiment

Angradi (1999), Zweig and Rabeni (2001)b

Reduced species density and diversity

D

Water filtration facility

Erman and Ligon (1988)

Reduced richness

D

Agriculture

Lemly (1982)b

Reduced taxon and EPT richness

D

Induced in field experiment

Matthaei et al. (2006)c

Altered composition and reduced EPT richness (patch scale)

D

Bank erosion

Larsen et al. (2009)a

Reduced EPT richness

D

Agricultural

Kaller and Hartman (2004)b

Altered functional composition/reduced abundance of filterers

D

Induced in field experiment

Bo et al. (2007)f

Impaired filter-feeding and reduced metabolic rate of mussels

S

Induced in experiment

Aldridge et al. (1987)b

Altered community structure/reduced density/ increased drift

S and D

Induced in experiment

Rosenberg and Wiens (1978); Larsen and Ormerod (2010)d,a

Altered community structure

D

Road construction

Extence (1978)a

Altered community structure

S and D

Agriculture

Richards et al. (1993)a

Decreased density of functional feeding groups and increased density of gathers.

D

Natural

Rabení et al. (2005)b

Altered richness, abundance, community composition, trait diversity and trait composition

D

Induced in field experiment

Larsen et al. (2011)a

Altered species/EPT density, richness and abundance

D

Altered land use

Niyogi et al. (2007)c

Reduced egg hatching

D

Induced in experiment

Kefford et al. (2010)e

Increased drift

D and Sa

Induced in field experiment

Culp et al. (1986)

Increased Baetis mayflies drift

D

Primary producers

Macroinvertebrates

Gibbins et al. (2005)a

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Table 4.1. Continued Impact

Sediment type

Source

Reference(s)

Reduced abundance

D and S

Desilting operation

Doeg and Koehn (1994)e

Reduced abundance

D

Agriculture

Berkman and Rabeni (1987)b

Decline in salmonid spawning habitat quality

D

Natural

Carling and McCahon (1987)b

Decline in salmonid spawning habitat quality

D

Coal mining

Turnpenny and Williams (1980), Lisle (1989)a,b

Decline in salmonid spawning habitat quality

D

Impoundment

Sear (1993)a

Reduced survival of salmonid eggs

D

Water filtration facility

Erman and Ligon (1988)b

Reduced survival of salmonid eggs

D

Agriculture

Soulsby et al. (2001)a

Reduced survival of salmonid embryos

D

Laboratory and field experiments

Greig et al. (2005)a

Reduced alevin survival

D

Laboratory and field experiments

O’Connor and Andrew (1998)g

Reduced survival to pre-eyed, eyed and hatched stages

D

Induced in field experiment

Julien and Bergeron (2006)d

Increased mortality of sac fry

D

Placer gold mining

Reynolds et al. (1989)d

Decreased growth and survival of juveniles

D

Induced in field experiment

Suttle et al. (2004)b

Reduced length and mass

D

Induced in field experiment

Shaw and Richardson (2001)d

Fish

a

Country of study: UK. Country of study: USA.

b c

Country of study: New Zealand. Country of study: Canada.

d e

Country of study: Australia.

Country of study: Italy.

f

Country of study: Ireland.

g

D, deposited sediment; D and Sa, deposited and saltating sediment; EPT, index named after the orders Ephemeroptera, Plecoptera and Trichoptera; S, suspended sediment.

Table 4.2. The effects of varying the concentrations of and the duration of exposure to suspended sediment on periphyton and macrophytes (from Bilotta and Brazier, 2008) Organism

SSC (mg/L)

Exposure time (hours)

Effect on organism

Reference

Macrophytes and algae

8



3–13% reduction in primary production

Lloyd et al. (1987)b

40



13–50% reduction in primary production

Lloyd et al. (1987)b

200



50% reduction in primary production

Van Nieuwenhuyse and LaPerriere (1986)b

2100



No primary production

Van Nieuwenhuyse and LaPerriere (1986)b

Phytoplankton

10

1344

40% reduction in algal biomass

Quinn et al. (1992)c

Aquatic moss

100

504

Extensive abrasion of leaves

Lewis (1973)a

500

168

Severe abrasion of leaves

Lewis (1973)a

100



Enhanced growth and filament length (low-flow velocities)

Birkett et al. (2007)b

200



Significant reduction in biomass and filament length

Birkett et al. (2007)b

0–6500



Abrasive damage and reduced biomass

Francoeur and Biggs (2006)c

Periphyton

a

Country of study: UK. Country of study: USA.

b c

Country of study: New Zealand.

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Table 4.3. The effects of varying the concentrations of and the duration of exposure to suspended sediment on macroinvertebrates (from Bilotta and Brazier, 2008, Collins et al., 2011, Jones et al., 2011)

a

Organism

SSC

Exposure time (hours)

Effect on organism

Reference

Cladocera and Copepoda

300–500 mg/L

72

Clogging of gills and gut

Alabaster and Lloyd (1982)g

Diptera

> 50 mg/L



Feeding inhibition

Kurtak (1978), Gaugler and Molloy (1980)b

Plecoptera/Trichoptera

1.5 mg/L

336

Feeding inhibition

Hornig and Brusven (1986)b

Bivalvia

600 mg/L

Intermittent exposure

Feeding inhibition/reduced metabolism

Aldridge et al. (1987)b

Cladocera

82–392 mg/L

72

Reduced reproduction and survival rates

Robertson (1957)b

Chironomids

300 mg/L

2016

90% reduction in population size

Gray and Ward (1982)b

Benthic invertebrates

62 mg/L

2400

77% reduction in population size

Wagener and LaPerriere (1985)b

Benthic invertebrates

743 mg/L

2400

85% reduction in population size

Wagener and LaPerriere (1985)b

Invertebrates

Pulses

456

Reduced abundance and richness

Shaw and Richardson (2001)d

Invertebrates

550–700 kg/50 m stream reach

840

Reduced richness, increased drift

Matthaei et al. (2006)c

Invertebrates

0.6–1.8 kg/m2

456

Reduced overall abundance and trait diversity

Larsen et al. (2011)a

Ephemeroptera, Plecoptera, Trichoptera

7.3–12.3%



Reduced relative abundance

Bryce et al. (2010)b

Stream invertebrates

130 mg/L

8760

40% reduction in species diversity

Nuttall and Bielby (1973)a

Benthic invertebrates

8 mg/L

2.5

Increased rate of drift

Rosenberg and Wiens (1978)d

Invertebrates

8–177 mg/L

1344

26% reduction in invertebrate drift

Quinn et al. (1992)c

Macroinvertebrates

133 mg/L

1.5

Drift increased (by 7-fold)

Doeg and Milledge (1991)e

Ephemeroptera

2680 mg/L



Increased drift

Ciborowski et al. (1977)d

Amphipoda Trichoptera

> 2000 mg/L

Varying exposure times

Drift and survival unaffected

Molinos and Donohue (2009)f

Mayfly (leptophlebiid)

1000 NTU

336

Mortality unaffected

Suren et al. (2005)c

Odonata

1000–1500 NTU

1

Reduced feeding efficiency

Kefford et al. (2010)e

1000 NTU

1

Increased survival

Kefford et al. (2010)e

Country of study: UK. Country of study: USA.

b c

Country of study: New Zealand. Country of study: Canada.

d e f

Country of study: Australia.

Country of study: Ireland. Country of study: Germany.

g

NTU, nephelometric turbidity units.

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Table 4.4. The effects of varying the concentrations of and the duration of exposure to sediment on fish (from Bilotta and Brazier, 2008, Collins et al., 2011) Organism

Sediment concentration (mg/L)

Exposure time (hours)

Effect on organism

Reference

Atlantic salmon

20



Increased foraging activity

Robertson et al., 2007c

Atlantic salmon

60–180



Avoidance behaviour/increased foraging activity

Robertson et al., 2007c

Chinook salmon

488

96

50% mortality of smolts

Stober et al., 1981b

Chinook salmon

207,000

1

100% mortality of juveniles

Newcomb and Flagg, 1983b

Sockeye and Coho salmon

800–47,000



80% reduction in successful egg fertilisation at sediment concentration of > 9000 mg/L

Galbraith et al., 2006bc

Coho salmon

2000–3000

192

Reduced feeding efficiency and immunity

Redding et al., 1987b

Coho salmon

400,000

96

Physical damage to gills; stress response

Lake and Hinch, 1999bc

Rainbow trout

47

1152

10% mortality of incubating eggs

Slaney et al., 1977c

Rainbow trout

Pulses

456

Reduced growth

Shaw and Richardson, 2001c

Brown trout

5838

8670

85% reduction in population size

Herbert and Merkins, 1961a

Arctic grayling

25

24

6% mortality of sac fry

Reynolds et al., 1989c

Arctic grayling

65

24

15% mortality of sac fry

Reynolds et al., 1989c

Arctic grayling

185

72

41% mortality of sac fry

Reynolds et al., 1989c

Country of study: UK.

a

Country of study: USA.

b c

Country of study: Canada.

Birkett et al., 2007), species diversity and organic content (Cline et al., 1982; Graham, 1990). In addition, an increase in flow and sediment levels can cause an increase in abrasion (Lewis, 1973; Francoeur and Biggs, 2006) and scouring (Steinman and McIntire, 1990); this reduces the attractiveness of periphyton to algal grazers (Graham, 1990).

macrophytes also leads to increases in channel roughness and water depth (Hearne and Armitage, 1993), thereby creating habitat diversity (Armitage, 1995). However, any decrease in macrophyte abundance can destabilise sediments leading to sediment resuspension (Madsen et al., 2001).

4.3 Macroinvertebrates

4.2.2 Macrophytes

Macroinvertebrates have the capacity to withstand occasional increases in deposited, saltating and suspended sediment levels (Ryan, 1991). However, increases in the levels of anthropogenically derived sediment reduce species abundance and diversity, and alter community structure. Moreover, as macroinvertebrates are a significant food source for fish, any changes in species abundance, diversity or quality will directly impact on fish populations. The various mechanisms of sediment impact on macroinvertebrates are illustrated in Figure 4.2.

Macrophytes are easily abraded by suspended sediments, but their presence can actually enhance the deposition and accretion of fine sediments (Wood and Armitage, 1997), which may ultimately lead to their elimination if substantial sediment accumulation occurs (Nuttall and Bielby, 1973). Macrophyte beds can reduce flow rates, thereby leading to an increase in sediment accumulation and a decrease in turbidity; this, in turn, leads to an increase in light availability and enhanced macrophyte growth (Madsen et al., 2001). The seasonal growth of in-stream and marginal

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Figure 4.2. Schematic showing the mechanisms by which macroinvertebrates are affected (directly and indirectly) by suspended, deposited and saltating sediment particles (modified from Jones et al., 2011).

4.3.1

Mechanisms by which deposited sediment affect macroinvertebrates

threshold level for EPT diversity in excess of 0.8–0.9% fine sediment (< 0.25 mm) substrate accumulation.

Studies have shown that deposited sediment can affect invertebrates as a result of burial (Wood et al., 2001, 2005), altered benthic substrate composition, interstical infilling, bed instability (Kaufmann et al., 2009), reduced refugia availability (Lancaster and Hildrew, 1993) and oxygen depletion (Swan and Palmer, 2000; Stead et al., 2004), which restricts the depth to which some invertebrates can penetrate deposited sediment. Sedentary species, such as the freshwater pearl mussel Margaritifera margaritifera (L), are particularly susceptible to deposited sediment accumulation that limits interstitial sediment oxygen supply (Geist and Auerswald, 2007; Osterling et al., 2008). Kaller and Hartman (2004) suggested a

4.3.2

Mechanisms by which suspended sediment affect macroinvertebrates

Suspended sediment affects macroinvertebrates by clogging (Alabaster and Lloyd, 1982) and abrasion of feeding mechanisms (Kurtak, 1978; Gaugler and Molloy, 1980; Hornig and Brusven, 1986; Aldridge et al., 1987); reduced feeding leads to reduced reproduction (Robertson, 1957), population size (Gray and Ward, 1982; Wagener and LaPerriere, 1985), and species abundance, richness (Shaw and Richardson, 2001) and diversity (Nuttall and Bielby, 1973). Suspended sediment reduces light penetration, thereby affecting invertebrate food sources, habitats

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and visual predators. Severe sediment loading may increase mortality rates (Hynes, 1970).

which sediment affects fish are described by Kemp et al. (2011).

Macroinvertebrates can utilise behavioural responses to protect delicate structures (Kurtak, 1978); for example, they may retract filter combs or switch food source [e.g. Brachycentrus (Trichoptera) switches from trapping food particles with filtering limbs (Gallepp, 1974) to grazing under conditions of high SSCs (Voelz and Ward, 1992)]. However, these behavioural responses may have implications for feeding and growth rates.

4.3.3

4.4.1

Effects of deposited sediment on salmonids

A decline in the quality of spawning habitats because of deposited fine sediment infiltration (Turnpenny and Williams, 1980; Carling and McCahon, 1987; Sear, 1993) has repercussions for egg survival and subsequent fish recruitment (Wood and Armitage, 1997). The deposition of fine sediment within redds, in the areas in which they infiltrate the gravel voids, reduces gravel permeability and porosity, and thereby adversely affects egg and alevin survival through diminished oxygen supply (Turnpenny and Williams, 1980; Erman and Ligon, 1988; Reynolds et al., 1989; O’Connor and Andrew, 1998; Argent and Flebbe, 1999; Shaw and Richardson, 2001; Soulsby et al., 2001; Armstrong et al., 2003; Greig et al., 2005; Julien and Bergeron, 2006; Heywood and Walling, 2007). The presence of sediment-associated oxygen-consuming material (Greig et al., 2005; Dumas et al., 2007) and/ or the inadequate removal of metabolic waste from incubating eggs can also be an issue. Conversely, coarser particles, which form a seal in the upper redd layer, can reduce fry emergence rates (Crisp, 1993). Fry and older fish may be affected by habitat alterations, such as a reduction/loss of juvenile rearing habitats and reduced water depths within pool areas, which are essential for adult salmonids, as a result of infilling due to the deposition of fine sediment (Waters, 1995).

Combined effects of suspended and deposited sediment on macroinvertebrates

Invertebrate drift is a natural process in fluvial systems, which expedites the movement of invertebrates among patches. Whilst moving particles may contribute to the dislodgement of invertebrates (Culp et al., 1983), the behavioural responses exhibited by some species suggest that increased drift may be used by invertebrates to evade the negative impacts (e.g. burial and clogging) of suspended, saltating or deposited particles (Ciborowski et al., 1977; Culp et al., 1986; Gibbins et al., 2007, Molinos and Donohue, 2009). An increase in invertebrate drift can reduce the benthic density and richness (Culp et al., 1983; Doeg and Milledge, 1991; Suren and Jowett, 2001; Larsen and Ormerod, 2010) of some sediment-sensitive taxa, while the abundance of sediment-tolerant taxa may be unaffected or, indeed, may increase (Kreutzweiser et al., 2005).

4.4.2

4.4 Fish Extensive research has been undertaken into the impact of fine sediment on fish (Turnpenny and Williams, 1980; Berkman and Rabeni, 1987; Carling and McCahon, 1987; Erman and Ligon, 1988; Lisle, 1989; Reynolds et al., 1989; Sear, 1993; Doeg and Koehn, 1994; O’Connor and Andrew, 1998; Shaw and Richardson, 2001; Soulsby et al., 2001; Suttle et al., 2004; Greig et al., 2005; Julien and Bergeron, 2006), which reflects their economic importance (Ryan, 1991). Most fish can tolerate short-term naturally occurring increases in sediment levels; however, some species, such as salmonids, are highly sensitivity to sediment (Waters, 1995). The main mechanisms by

Effects of suspended sediment on salmonids

The effects of suspended sediment on salmonids can be categorised as behavioural, sub-lethal or lethal, depending on the SSC and the exposure time. Behavioural responses range from an alarm reaction, the abandonment of cover to the avoidance of sediment fluxes (Servizi and Martens, 1991). Sub-lethal effects are more wide ranging and include a reduction in the tolerance to toxicants (Lloyd et al., 1987), a decrease in disease resistance (Redding et al., 1987), damaged gills (Herbert and Merkins, 1961; Redding et al., 1987), the interruption of gas exchange and osmoregulation (Bruton, 1985; Waters, 1995), the disruption of development (Suttle et al.,

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2004) and a delay in growth (Shaw and Richardson, 2001; Sutherland and Meyer, 2007). Lethal effects due to long-term exposure to high SSCs can lead to a population-level response resulting in an increase in mortality rates (Stober et al., 1981) and the loss of species from affected reaches (Birtwell et al., 1984).

Interactions between sediment sources and associated contaminants, the sediment concentration, the duration of exposure to sediment, and sediment particle size and chemical composition, along with the sensitivity of the particular taxa, certain life-history traits and species life-stage are all important factors that determine the ecological effects of sediment (Swietlik et al., 2003; Bilotta and Brazier, 2008; Collins et al., 2011).

sustained impacts on biota (Collins et al., 2011). A meta-analysis study by Newcombe and McDonald (1991) found that the ranked response of aquatic biota was poorly correlated with SSC and more strongly correlated with sediment intensity (the product of sediment concentration and the duration of exposure). This suggests that both the concentration and the duration of exposure need to be assessed in order to predict the effects of sediment on aquatic biota. Studies on the effects of sediment on fish have shown increases in foraging activity and avoidance behaviour (Robertson et al., 2007), gill damage (Lake and Hinch, 1999), effects on the mortality of eggs, fry and smolts (Slaney et al., 1977; Stober et al., 1981; Reynolds et al., 1989), and reductions in feeding efficiency (Redding et al., 1987), egg fertilisation (Galbraith et al., 2006) and growth rates (Shaw and Richardson, 2001) depending on sediment concentration and the duration of exposure.

4.4.4

4.4.6

4.4.3

Sediment characteristics that influence ecological effects

Particle size and geochemical composition

As previously stated, fine sediment is a natural and vital component of aquatic ecosystems (Wood and Armitage, 1997; Owens et al., 2005) that ensures habitat heterogeneity and ecosystem functioning (Yarnell et al., 2006). However, anthropogenically derived fine sediment fluxes can alter temporal and spatial fluvial sediment dynamics significantly, and increase the delivery and transport of sorbed contaminants, including nutrients, heavy metals, pesticides and pathogens, with the < 63-µm sediment fraction to aquatic systems (Salomons and Forstner, 1984; Stone and Droppo, 1994; Owens et al., 2001, 2005).

Sediment geochemistry, which is influenced by soil type and catchment geology, and its physical characteristics, including the size, shape and angularity of particles, will also influence ecological impacts. For a given turbulence condition and sediment type with identical particle densities and shapes, smaller particles usually stay in suspension for longer than larger particles (Schindl et al., 2005). Under low-flow conditions, fine particles remain close to or at the water column surface (Schindl et al., 2005); this affects the respiratory and feeding mechanisms of invertebrates and fish. Coarse particles settle out of the water column (Schindl et al., 2005), which potentially affects salmonid redds and invertebrate habitats, as previously described (Greig et al., 2005).

4.4.5

Sediment and associated contaminants

The concomitant effects of the sedimentation and nutrient enrichment that are associated with managed and cultivated catchments may result in additive, antagonistic or synergistic biological responses as a result of complex interactions between these stressors (Lemly, 1982; Townsend et al., 2008; Matthaei et al., 2010). Some studies suggest that temporal patterns of sediment and nutrient disturbances drive ecosystem response to multiple stressors (Molinos and Donohue, 2010), whist others have shown that localised catchment modifications due to “poaching” and/or bank erosion may result in reach- and patch-scale ecological impacts (Larsen et al., 2009).

Concentration and duration of exposure

Whilst extremely low concentrations of sediment can negatively affect aquatic species, the duration of exposure (Tables 4.2–4.4) is also a critical factor (Newcombe and McDonald, 1991; Newcombe and Jensen, 1996; Swietlik et al., 2003; Cocchiglia et al., 2012b). Short-term exposure may result in ephemeral effects, whereas long-term exposure will result in

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5

Measuring and Monitoring Suspended Sediment Concentrations and Loads

5.1 Introduction

section, although all approaches benefit from an initial reconnaissance study, as summarised below.

It has long been known that it is difficult to estimate fluvial SSLs (“silt fluxes”) accurately and precisely (see, for example, Walling and Webb, 1981a; Ferguson, 1987). At the simplest level, it requires the determination of river discharge (Q) and spatially and temporally representative SSCs. Thus, the instantaneous SSL in kg/s, is given by: SSL = Q × SSC

5.1.1

The value of reconnaissance suspended sediment data

Manually collected, reconnaissance SSC or turbidity data – especially from repeat “storm-chasing” runs during elevated SSCs – can give vital information, at an early stage, on key indicators of the sediment transport system, and can help with planning a more spatially and temporally intensive sampling and monitoring campaign; this is vital for the accurate and precise estimation of suspended sediment fluxes. Key indicators include:

(Equation 5.1)

In Equation 5.1, Q is the water discharge in m3/s and SSC is the instantaneous suspended sediment concentration in g/L (i.e. kg/m3). The estimation of sediment fluxes over a given period requires the integration of this simple equation over time, and this is discussed in Section 5.9. River discharge measurement techniques are well reviewed elsewhere (see, for example, Hershey, 1999). However, the robust estimation of fluvial SSCs is far from simple. Therefore, this section focuses on the methodologies, techniques and instrumentation available for the estimation of SSC in rivers.

1. approximate fluvial SSCs; 2. suspended sediment particle size distribution (this can give an indication of the sediment sources, and help to predict likely cross-sectional variations in SSC); along with information on fluvial SSCs, this information can help to guide the selection of the most appropriate turbidity meters, which can be geared towards specific ranges of SSCs and particle size distributions (e.g. via their beam intensity or the wavelength of the instrument);

Ideally, the best way to obtain the most accurate estimates of SSC is to sample river waters by hand at very high resolution in (1) time, in order to capture variations related to hydrograph events (e.g. take samples at 15-min intervals before, during and after storms); and (2) space, that is, take samples from water surface to river bed, in order to embrace the entire water column, and across the full channel width to account for any cross-sectional variation in SSC, which is especially pronounced in large rivers with low turbulence and coarse suspended sediment particle size distribution. Repeated field sampling missions should then be followed by the laborious processing of thousands of water samples in a laboratory to determine SSCs. However, although this might be possible for short-term studies, for longterm investigations of fluxes, the resources required for this highly labour-intensive approach would be impracticable. Thus, various compromises have to be made; these problems, solutions and alternative approaches and methodologies are outlined in this

3. suspended sediment composition, which can indicate likely sediment sources; 4. the extent of any width or depth variation in SSC across the gauging cross-section; 5. catchment-scale and downstream changes in SSC, which can provide clues on the locations of sediment sources, storage and conveyance losses, and can inform decision-making with regard to the choice of sites for monitoring stations; 6. Q data; if such data are available or estimable, preliminary instantaneous SSL data can be obtained using Equation 5.1 and used as guidance for later turbidity meter and automatic water sampler installation, siting and intake, and instrument head locations (e.g. for channel-edge sampling).

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Reconnaissance runs throughout the catchment more widely (not just of the stream network itself) during rainstorms can also help to identify key sediment sources and delivery processes.

5.2

Manual Sampling

5.2.1

Cross-sectional variations in SSC

Doppler current profiler (ADCP) data have recently helped to define spatial patterns in cross-sections and, for flux investigations, SSC could ideally be representative of the whole cross-section, but this is rarely possible to achieve with high temporal resolution monitoring. Normally, just a single, fixed or floating, intake or monitoring point is placed at the channel edge. Therefore, especially in large rivers of low turbulence and/or with coarse suspended sediment in which strong lateral and vertical SSC gradients develop, the following three additional procedures can be introduced to help attain the most accurate and representative values of suspended sediment flux:

It has long been known that significant cross-sectional variations in SSCs exist (e.g. Guy and Norman, 1970; Horowitz et al., 1989): SSCs generally increase towards the bed, especially for coarser particles, and towards the channel centre (Figure 5.1 and Figure 5.2). Figure 5.3 shows data supporting this, and although datasets are rare, channel centre SSCs can be twice that of those at the channel edge in some large rivers (Nordin and Richardson, 1971). Acoustic

1. Attempts should be made to achieve a spatially representative position for the intake or monitoring point (e.g. not in a turbid “hot-spot”).

v

cs

v cs

unmeasured strip

unmeasured strip

0

0

0 Suspended Sediment Concentration

Suspended Sediment Flux

Velocity of Water flow

Depth-integrating sampler

Figure 5.1. Schematic of SSC cross-sectional variations.

Figure 5.2. Schematic cross-sectional variation in flow velocity, SSC and sediment flux.

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Water surface

9

7

1

> 0.26 mm

2

0.125 – 0.25 mm

3

0.62 – 0.125 mm

4

0.016 – 0.062 mm

5

0.002 – 0.016 mm

6 < 0.002 mm

Distance above streambed (m)

8

0 Scale of suspended sediment concentration (parts/million) 0

100

200

300

50

100

150

Water velocity (cm/sec)

Figure 5.3. Vertical distribution of concentration of various particle sizes in a stream section (after Nordin and Richardson, 1971). 2. SSCs at this single point should be compared with those from occasional depth-integrated manual sampling points across the channel width.

around the sampler. This is called “isokinetic sampling” and is very important for representative suspended particulate sampling.

3. If the single-point concentration value tends to significantly over- or underestimate crosssectional average concentrations, a suitable correction/calibration factor should be determined and applied.

Isokinetic sampling is described in Glickman (2000) as “any technique for collecting airborne (waterborne) particulate matter in which the collector is so designed that the air stream (water stream) entering it has a velocity equal to that of the air (water) passing around and outside the collector. The advantage of isokinetic sampling consists in its freedom from the uncertainties due to selective collection of the larger, less easily deflected particulates. In principle, an isokinetic sampling device has a collection efficiency of unity for all sizes of particulates in the sampled air or water”.

5.2.2. Sampling to detect cross-sectional variation in suspended sediment For very turbulent flows, or for fine suspended sediment size distributions, lateral and vertical variability in SSC will normally be less notable than cross-sectional variation. Sediment mixing is also likely to vary with stage for a given cross-section, although few datasets demonstrate this. To check for sediment mixing, spatial sampling should be done with specially designed point- or depth-integrated isokinetic samplers (Guy and Norman, 1970; Davis, 2005 in Sabol et al., 2010), as pioneered by the United States Geological Survey (USGS).

Isokinetic sampling

At each vertical, such samplers need to be lowered at an even rate to the bed and raised again to the surface. This allows sampling of most of the water column, except the lowest few centimetres (e.g. the lowest 9 cm for the US DH-48 sampler). Transits through the water column are repeated, if necessary, at the same vertical until the bottle is approximately 70% full, thus avoiding overfilling, which can bias the sample. This is called the “equal transit rate” method (Guy and Norman, 1970; Edwards and Glysson, 1999; Gray and Gartner, 2009).

The key feature of a USGS isokinetic sampler is that the bottle is allowed to fill through only a carefully machined and chamfered 6-mm special nozzle that, if pointed upstream, allows water to flow into the sampler at the same velocity (±10%) as water passing

This vertical sampling procedure is then repeated at “no less than 3 and usually 5 to 10 verticals in a given cross section” (Dunne and Leopold, 1978) to capture the lateral variations in SSC (in a similar way as is done for stream gauging using impeller

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or electromagnetic current meters). Point samplers can be used to sample at only given depths for each vertical.

SSCs (as a spin-off from the normal usage of ADCP for velocity profiling), are outlined by Gray and Gartner (2009). Wall et al. (2006) describe the typical procedures used to estimate fluvial and tidal SSCs from ADCP measurements. They stress that, although ADCPs were not designed for SSC measurements, they can be useful, but only after considerable investment in time and appropriate software to obtain a suitable transfer function between the acoustic signal and SSC (see also Filizola and Guyot, 2004). A particular strength of the technique is its ability to produce a “snapshot” of cross-sectional variation in SSC, especially for larger rivers.

In shallow rivers, depth-integrating samplers can be fitted to a wading rod (e.g. US DH-48 sampler developed by the USGS) and hand-lowered to the bed by a fieldworker in waders or a drysuit. In deep rivers, depth-integrated sampling will need to be carried out from boats or at bridges by deploying the sampler on a long rod or cable. Samples are transferred into plastic sample bottles, sealed and labelled, and transported back to the laboratory for analysis of SSC. To do this, samples of known volume are simply filtered normally through 0.45-μm filter membranes. Vacuum filtration can be used if clogging of the filter paper/membrane occurs. SSC can be calculated as follows: SSC = wt/vol

5.4

Many authors have argued that suspended sediment data were traditionally collected using a range of manual sampling devices and that “the associated logistical problems and financial constraints mean that most manual sampling strategies fail to coincide with the main periods of sediment transport, i.e., flood events” (Collins and Walling, 2004). Such sampling programmes, if used alone, significantly underestimate actual suspended sediment fluxes and yields. Thus, although manual sampling can provide very useful data on spatial variations in SSC across, and between, single cross-sections and catchments, it is rarely sufficient to provide the necessary high-resolution time series of SSC, which are absolutely crucial for accurate and precise estimates of sediment fluxes and SSC.

(Equation 5.2)

In Equation 5.2, “wt” is the dry weight of sediment (in g) and “vol” is the water sample volume (in L). (Note that organic material may need to be removed to determine the minerogenic component of the suspended material.)

5.2.3

Time-integrating samplers

To collect larger samples of suspended sediment for chemical analysis (e.g. for source tracing), timeintegrating samplers are available. These are tubes with intake and outflow nozzles at the upstream and downstream ends, respectively. These samplers can be left in the flow for some time to collect sediment, which settles in the sampler (Figure 5.4). A useful video showing the use of these samplers is available on YouTube (http://www.youtube.com/ watch?v=PnSm4hNAJ4Q).

The deployment of automatic water samplers (AWSs) is one way to address this need to sample at high temporal resolution. AWSs can be placed next to rivers in waterproof and vandal-proof housings, and set up to automatically extract water samples at, for example, 30-minute, 60-minute or 3-hour intervals (depending on the aims and resources of the study, and the flashiness of the river flow regime) and store the water samples in a crate on site. Crates are replaced at intervals, and the samples are retrieved and transported to the laboratory for processing and filtration. AWSs can be seen in action on YouTube (at: http://www.youtube.com/watch?v=kQl1qk7z8do).

Infiltration baskets can also be inserted into river beds to measure the accumulation of fine sediment. The mass and composition of the fine sediment can be quantified upon removal of the basket. An example from Heywood and Walling (2007) is shown in Figure 5.5.

5.3

Automatic Sampling for Suspended Sediment Concentration Time Series

Acoustic Doppler Current Profiler Method

However, the deployment of “passive” AWSs is likely to generate numerous samples at low flow and low SSC,

The principles of the ADCP (essentially acoustic backscatter) method, which is used to estimate

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typically 1 m long

4 mm dia

4 mm dia

typically 10 cm diameter

Bed of channel

Figure 5.4. Time-integrating suspended sediment sampler for collecting large amounts of suspended sediment for composition analysis (after Collins et al., 2010).

Figure 5.5. Infiltration basket for capturing fine sediment in gravel river beds (photo: Liz Conroy).

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and very few samples at high flow when suspended sediment fluxes peak. In many catchments, 95% of the sediment flux occurs only 10% of the time, normally during high-flow events. Therefore, it is important to use active, programmable AWSs (such as the ISCO, American Sigma or EPIC models), which can, if desired, lie dormant during periods of low flow/turbidity when sediment transport is minimal, and activate only if linked stage and/or turbidity triggers are exceeded. Such event-actuated sampling avoids the unnecessary use of valuable crated bottles, and saves them for the vital samples taken during key sediment transport events, which are fundamental to accurate sediment flux estimation. Of course, for completeness, some low-turbidity samples are desirable.

estimates of SSC can be obtained. Turbidity metres can ease these problems considerably, and are discussed below.

5.5

Turbidimetric Instrumentation

5.5.1 Introduction Turbidity “is an expression of the optical property of a medium which causes light to be scattered and absorbed rather than transmitted in straight lines through the sample” (American Public Health Association et al., 1995). Turbidimetry is a surrogate method for estimating SSC. However, the field is hampered by a lack of consistency with regard to units, sensors and calibration techniques, and there are many papers on these issues (e.g. Lawler, 2005a; Lawler, 2015). Turbidity used to be monitored manually by lowering a Secchi disk into the water, and simply noting the depth at which the disk markings became invisible to an observer on the surface (usually in a boat); although this is a subjective method, it provided the crucial long-running dataset (Figure 5.6) that was used to address the well-publicised problems of increasing turbidity over the last few decades in Lake Tahoe on the California/Nevada border (Jassby et al., 1999), which led to legal disputes with loggers in an attempt to prevent the loss of amenity resources in a key tourist area.

The intakes for AWSs are normally fixed at a single point in the cross-section near the bank (below the water level at all times, but above the level at which fine sediment clogging or damage by bedload transport may occur); however, floating intakes can also be used for sampling at a consistent proportional depth. Although AWSs can provide strong datasets on SSCs and therefore SSLs, they have three disadvantages: (1) a relatively low frequency of sampling, typically limited by crates of no more than 24 bottles; (2) the lack of available capacity for sampling sequences of floods, when the bottles have already been filled by sampling from the initial event; and (3) the generatation of large volumes of samples that have to be processed and filtered in the laboratory before

However, as reviewed by Gray and Gartner (2009) and Lewis (1996), many relevant and more modern techniques for measuring turbidity are now available.

Figure 5.6. Declining water clarity in Lake Tahoe, measured using the Secchi disk. The clarity decrease (turbidity increase) is thought to have been driven by fine sediment discharge from the surrounding lake catchment as a result of, for example, logging activities (after Jassby et al., 1999).

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The units of measurement are typically FTU (formazin turbidity units) or NTU (nephelometric turbidity units). One of the longest running investigations of turbidity was carried out in the USA by Jones and Schilling (2011); these authors demonstrated that sediment load decreased during the 20th century and recommended focusing management on reducing mobilisation.

5.5.2

rapidly changing SSCs – capturing which is beyond the capability of hand-sampling or even AWSs – much contemporary research on suspended sediment dynamics and fluxes is based on the automatic and high-resolution monitoring of turbidity as a surrogate variable for SSC (see, for example, Lawler, 2005a; Walling, 1974). The great advantage of measuring water turbidity is that it can now be monitored automatically using electronic turbidity meters, such as the Partech IR40 and IR15 instruments, which are connected to data loggers. Data loggers (e.g. those manufactured by Campbell Scientific Ltd) can be programmed to simply scan and store turbidity data at high resolution (e.g. 15-minute frequency), or to store, if desired, the average of several scans if logger memory capacity is limited. Hence, it can much more usefully capture changing suspended sediment transport, including peaks, troughs and flat-lining in the SSC signal. These attributes are essential for the robust quantification of sediment concentrations and hence fluxes; in addition, the time-dependent sediment behaviour with respect to other events can aid understanding of sediment dynamics and sources.

The turbidimetry principle

The principle of turbidimetry is essentially simple: a beam, normally in the infra-red (IR) spectrum, of constant and known initial intensity, is emitted from a very high specification light-emitting diode (the kind used on aircraft control panels), and transmitted to an appropriate IR sensor across a known gap through which river water is allowed to pass. Particles within the gap, such as fine sediment, scatter or absorb some of the photons in the beam. This attenuates the beam intensity, as recorded by the sensor in the instrument. This reduction in the IR light recorded by the sensor is directly related to the concentration of suspended matter in the light path. It is well known, however, that transmittance and scatter are functions not just of SSC, but also the “number, size, colour, index of refraction, and shape of suspended particles” (Gray and Gartner, 2009). Particle size distribution is a key issue: a given mass (or concentration) of fine sediment will cause greater light scattering (and therefore beam attenuation) than the same mass of coarser sediment. Fine sediment has a greater specific surface area than the same mass of coarse sediment, and the particle specific surface area and roughness largely control the light scattering ability of sediment (Ward and Chikwanha 1980; Lawler, 2005a). Indeed, data from Ward and Chikwanha (1980) show that, for the same SSC (1000 mg/L), particles in the 12- to 18-μm diameter range absorb an order of magnitude more light than particles in the 30- to 50-μm diameter range. Thus, if the particle size distribution of the SSL varies in space and time, this will affect the applicability of the calibration relationship.

5.5.3

Turbidity meters can be mounted directly in the river (normally at the channel edge), although they can be vulnerable to damage from large floating objects, clogging by organic and other debris, or bio-fouling of the optics (e.g. with algae). Therefore, such instruments require periodic (e.g. weekly) cleaning in order to preserve the integrity of measurements. Some turbidity meter models are self-cleaning, with a motorised optics wiper system powered by an on-site battery or solar panel. To obviate some of the difficulties of in-stream deployment, turbidity instrumentation can instead be installed in bankside housings away from potential sources of mechanical damage and vegetation fouling. In such cases, river water can be automatically withdrawn by pump or vacuum and directed to the instrument in this “safe house” (e.g. England and Wales Environment Agency automatic water quality monitoring stations; Lawler et al., 2006). Turbidity meters work especially well at sites in which little seasonal or storm-event change in suspended sediment composition takes place; this should be assessed in any detailed investigation. However, any uncertainties regarding the role of such confounding variables locked into the turbidimetry principle are

Turbidity meters

Turbidity meters can be used as hand-held portable devices for “snapshot” measurements and vital, early field reconnaissance. However, their real value lies in the automated monitoring of SSC. Because of typically

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usually outweighed by the huge advantages of the continuous high-resolution datasets obtained.

Conventional calibrations used to be made with formazin, which is a polymer that was developed in 1926. Although formazin is considered a carcinogen, it is still used as a primary standard today (Sadar, 1998), and can be used for calibration as an alternative to sediment from the river or catchment being monitored. An effective method is simply to match up SSCs from samples withdrawn by on-site AWSs with synchronous, usually automated, turbidity meter measurements made as close as possible to the sampler intake point. If this is not possible, then large quantities of turbid water can be taken from the river for laboratory calibration purposes. Another option is to mix up a stock solution of turbid water using fine sediment (e.g. in the 1- to 250-μm diameter range) derived from likely, suspendable sediment sources upstream of the intended monitoring site (e.g. from river banks, exposed bars, fine bed sediment in pools, field poaching sites, farm tracks or hillslope gullies). The first step of calibration involves measuring the turbidity of a stock solution with the maximum expected SSC (e.g. 1000 mg/L). This stock is then sequentially diluted by adding distilled water in known volumes to produce a range of turbidity values (e.g. 20 values) until the lowest turbidity value is reached. Associated turbidity meter readings are taken for each known SSC. These high-to-low turbidity measurements minimise noise in the calibration by avoiding the need to add fresh sediment – possibly of a different particle size distribution – to the solution.

Turbidity meters can also be deployed in the tidal zone (often a zone of repeated sediment deposition and re-erosion) to monitor the magnitude and migration of the estuarine turbidity maximum which, in turn, can partly be a function of fluvial sediment flux (e.g. Mitchell et al., 2003).

5.5.4

Laboratory and field calibration of turbidimetric instrumentation

Data are often reported directly in FTU or NTU. However, for many geomorphological or sediment transport investigations, there is a need to report results in terms of SSLs or fluxes. This entails calibration, which can be difficult. Generic calibration factors between turbidity and SSC do exist, and can be provided by instrument manufacturers. However, it is preferable to derive a site-specific calibration factor for turbidity based on known SSCs, because turbidity–SSC relationships typically vary widely from river to river, station to station and over time, largely in response to changes in the composition of the SSL (see, for example, the field calibrations illustrated in Figure 5.7 and Figure 5.8).

SSC (mg/L)

There are a number of possible approaches to calibration. Laboratory procedures to derive calibrations are given in Lawler and Brown (1992) and Old et al. (2003).

Turbidity (NTU) Figure 5.7. Turbidity versus SSC: calibration for the urbanised area at James Bridge, River Tame, Birmingham, UK (Lawler et al., 2006).

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Suspended sediment concentration (g/L)

12

10

8

6

4

2

0

0

1

3

2

4

5

6

7

Turbidity measurement (Volts) Figure 5.8. Turbidity versus SSC: calibration for the large Skaftá river, south Iceland (Old et al., 2005a).

Manufacturers of turbidity instrumentation (e.g. Hach or Partech) can supply generic calibration factors, stock calibration solutions and gel slides which have a fixed assemblage of particles embedded within them, for a range of sediment concentrations. Such gel slides can be placed between the turbidity meter beam source and the sensor.

litre of water (SSC = 1000 mg l–1) might produce an OBS signal of 1 Volt; whereas a gram of sand with a grain size of 100 microns would produce only a 0.25-Volt signal, with other factors such as shape and mineral composition being the same” (Downing, 2008b). This is illustrated for a range of particle size distributions in Figure 5.9 and Figure 5.10.

5.6

A typical OBS field monitoring procedure is helpfully described by Chappell et al. (2011, p. 94) as follows:

Optical Backscatter Sensor Instrumentation

At this station an OBS-3+SB-2.5-T4 turbidity probe (D & Instruments / Campbell Scientific Inc., Logan, USA) mounted on a self-cleaning Hydro-Wiper device (Zebra-Tech Ltd, Nelson, New Zealand) was attached to a protective steel manifold. A CR1000 data-logger (Campbell Scientific Inc., Logan, USA) was used to monitor turbidity every 30 seconds (on both 0–2000 and 0–4000 Nephelometric Turbidity Units or NTU ranges), with the average stored every 1 minute. Averaging and storage was changed to every 10 minutes after the first month of monitoring. A relationship between turbidity (NTU) and suspended sediment concentration (mg l–1) was established using water samples collected from the Ramu River over the ranges of 37 to 810 NTU and 74 to 2676 mg l–1.

The optical backscatter sensor (OBS) and nephelometer are similar to turbidity meters except that these instruments directly measure the amount of light or IR scattered by particles in the water column, rather than beam attenuation itself. For OBS instruments, backscatter is measured at 180° to the incident beam, that is, OBS instruments measure the IR radiation that is reflected back to the emitting sensor from particles approximately 0.5–30 cm from the front of the device, and outputs can be monitored automatically to create a useful time series of SSCs (e.g. Ruhl et al., 2001). As for turbidity meters, the particle size distribution of suspended sediment influences OBS instrument turbidity measurements. For example, “one gram of silt, with a grain size of 10 microns, suspended in a

43

SILTFLUX Literature Review

Light absorbance (ic)

1.0

< 5 µm 0.3

0.1

0.03

30

300

100

30-50 µm

12-18 µm

6 - 8 μm

3000

1000

10 ,000

Suspended sediment concentration (mg/L) Figure 5.9. Dependence of light absorbance on sediment particle size (after Lawler, 2005a). 10000

AY

SA

ND

CL 1000

500

SIL

T

OBS output (mV)

SIL

T

5000

100

10

50

100

500

1000

5000

10,000

50,000 100,000

Suspended sediment concentration (mg/L) Figure 5.10. Effect of particle size on OBS response (Downing, 2008b).

5.7 Laser In Situ Scattering and Transmissometry

logarithmically spaced size classes (Thonon and Van der Perk, 2003), within a total range of 1.5–250 μm (which is a typical fluvial suspended sediment range). Particle size distribution information may be used to pinpoint sediment source types or locations and pathways, and hence identify the key contributors to and explanations for any SSL problems in upstream catchments.

Laser in situ scattering and transmissometry (LISST) instrumentation also provides key information on sediment transport (e.g. Melis et al., 2003; Old et al., 2006; Gray and Gartner, 2009). LISST instruments are widely used by the USGS (e.g. Gray and Gartner, 2009) and are increasingly being used in Europe (e.g. Thonon and Van der Perk, 2003). In particular, LISST sensors usefully provide particle size distribution information for the suspended sediment in 32

Different LISST instrument models are available for different contexts (e.g. oceanic or fluvial) and various size classes. For example, the Sequoia model

44

D. Lawler et al. (2010-W-LS-4)

(see http://www.sequoiasci.com/) is a “streamlined isokinetic version of the LISST-100X developed for river sediment monitoring. The instrument senses the river current velocity using a pitot tube and adjusts (sampler) intake velocity to match. The river water is thus pumped through the instrument at the same speed as the water flowing around the instrument, making the sampling isokinetic”; this helps to ensure representative sampling. Data are transmitted from the submersible sensor head to a data logger on the bank.

1. The many non-linearities in the sediment transport system, especially with regard to the relationship between Q and SSC (Figure 5.11), which are often power functions (e.g. Figure 5.12), are problematic. 2. There are problems related to the temporal variability of river flows and the variable SSC for a given flow. Sediment rating curves show that SSC can vary by 2 to 3 orders of magnitude for a given flow [as in Figure 5.11 from Walling and Webb (1981a)]. This is quite different from a stage-discharge rating equation in which, in the absence of geometrical changes of the channel at the gauging station, a given stage is always associated with the same discharge (notwithstanding rising and falling limb water surface slope differences as the flood wave passes).

As with turbidity meters, LISST optics should be cleaned at approximately weekly intervals to remove biological fouling, algal growth etc., and more frequently under some conditions (e.g. tidal rivers).

5.8

Remote Sensing of Suspended Sediment Concentration

It is also possible to obtain useful data on the spatial variation of SSCs of surface and near-surface waters using remote sensing imagery from appropriate satellite, helicopter or fixed-wing airborne platforms (see, for example, Ruhl et al., 2001). Advanced very high resolution radiometers (AVHRR) from satellites have proved very useful in this regard (e.g. San Francisco Bay; see Ruhl et al., 2001). Such imagery is especially helpful for large-scale water bodies, rivers and estuaries because it can show, for example, turbidity hot spots and high-turbidity plume dynamics, especially in a relative sense. The imagery and AVHRR reflectance has to be calibrated against actual SSC values obtained using other methods in order to deliver usable data on SSCs (see, for example, Ruhl et al., 2001). This is done by synchronous sampling or monitoring at the same place and time of the image capture.

5.9

3. Most suspended sediment transport occurs over a very limited time. For example, Walling et al. (1992) found, for the River Exe at Thorverton in south-west England for the 1978–1980 period, that 50% of the SSL was present just 1% of the time, and 90% of the total SSL was present only 5% of the time. It is crucial, therefore, to sample or monitor during such key periods. Typically, this leads to huge scattering with regard to Q–SSC relationships (sediment rating curves). 4. There are characteristically log–linear relationships between Q and SSC. 5. Hysteresis effects drive different SSCs for a given flow (during storm events and different seasons). 6. Cross-sectional SSC variations may complicate single-point (e.g. channel-edge) sampling or monitoring, etc.

Estimation of Suspended Sediment Loads

7. Changing sediment particle size distribution effects, during and between storm events and seasons, can degrade simple relationships between turbidity and SSC.

5.9.1 Introduction As shown in Equation 5.1, SSLs are derived from the product of SSC and Q. Although apparently simple, the calculation of fluvial sediment fluxes is fraught with difficulty, and the process is often referred to as an “estimation”. Suspended sediment flux estimates need to maximise accuracy (i.e. proximity to the actual value) and precision (i.e. high repeatability). Problems include the following:

Ensuring that sampling/monitoring is quasi-continuous is key to estimating SSLs. Walling and Webb’s (1981a) pioneering study demonstrated the importance of this by using a re-sampling approach on data from the River Creedy in Devon, England. They first computed “true” SSLs from a master high-resolution dataset, obtained from samples collected at 60-minute intervals. They then compared these values with those

45

SILTFLUX Literature Review

5000 River Tyne at Bywell 1959-1961

Suspended sediment concentration (mg/L)

500

100 April Oct ober

50

1000

Suspended sediment concentrations (mg/L)

1000

Oct ober - April

10 5

River Exe at Thorverton (1971-1973) April - Sept

October - March

100

10

1 1 3

5

50

10

500

100

1000

1

Discharge (m3/s)

10 Discharge (m3/s)

100

Figure 5.11. Examples of SSC–Q relationships for two British rivers (after Walling and Webb, 1981b). 10,000 11a

5,000 4,000

Suspended sediment conc. (mg/L)

3,000 2,000

4a

17

11b

4b 14

5

9

1

8 13

1,000

10

7

3

500 400

15

300 17 200

11a 12

100

River / Stream

50 40

1 White Cart 3 Clyde 4a Tyne (summer) 4b Tyne (winter) 5 Rother 6 Slapton Wood Stream 7 Creedy 8 Exe 9 Dart 10 Ystwyth 11a Rosebarn summer 11b Rosebarn winter 12 Coachwater Drain 13 Wye 14 Severn 15 Usk 17 Crew

30 20

10 14___ 5 4

11b

3 2

1 6 17

9

3

1 0.001

5

7

0.002

13

15

10 4a

0.005

4b

8 0.01

0.02

0.05

0.1

0.2

0.5

1

2

5

Discharge (m3s-1km-2) Figure 5.12. Relationships between suspended sediment and area-weighted Q for several named British rivers (after Walling and Webb, 1981b).

46

D. Lawler et al. (2010-W-LS-4)

obtained from sampling at successively longer time intervals (e.g. daily, weekly, 2-weekly intervals) to mimic a low-resolution field investigation. Such studies demonstrate that very large load errors can arise with low temporal resolution sampling (see, for example, Walling and Webb, 1981a, 1985).

be sufficiently long to be representative of longer term sediment transport conditions, and include a range of hydrological and meteorological events, especially storm sequences during which most sediment transport occurs. Total suspended sediment load, (Ltot,) is then evaluated as the flux integral over a given period, as in Ferguson (1987):

Several methods have been evaluated rigorously by Walling and Webb (1981a, 1985), and further examinations of the statistical methodologies involved have been published by Ferguson (1987) and Clarke (1990a, b). Phillips et al. (1999) evaluated 22 SSL estimation equations for catchments in northern England.



T/δT

(Equation 5.3)

In Equation 5.3, in SI unit terms, Ltot is the total SSL over time (in kg/s), T is the time (in seconds) over which loads are calculated, δt is the fixed sampling or monitoring interval used (in seconds), Ci is the instantaneous SSC (in g/L or kg/m3) at time i, and Qi is the water discharge at time i (in m3/second).

However, the problems associated with SSL estimation persist and more recent examinations include those by Webb et al. (1997), who investigated British rivers; Phillips et al. (1999), who examined rivers in the UK and showed that errors are magnified in larger catchments; Horowitz (2003), who, on the basis of sediment rating curves for US and European rivers, recommended that sampling is done on a hydrological, rather than a calendar, basis to minimise errors; Crowder et al. (2007), who studied SSL–Q relationships for US Midwest basins; and Brown et al. (2010), who used multiple load estimation approaches for basins in the UK, Germany and the Netherlands. As Moatar and Meybeck (2005) argued, some of these issues also apply to nutrient load estimation: “concentration data are still commonly the limiting factor on the quality of river flux estimates”.

5.9.2

Ltot = ∑ i=1 Ci.Qi.δt

A number of variants of Equation 5.3 exist, depending on data (Q and SSC) availability and national conventions (see, for example, Horowitz, 2003).

5.9.3

Sediment rating curve method: extrapolation techniques

If automatic, continuous monitoring of turbidity is not practicable, then a reasonable, but less accurate and precise, alternative involves the use of sediment rating curves. These have been exhaustively discussed by numerous investigators from several disciplines (e.g. Walling and Webb, 1981a; Ferguson, 1986, 1987; Horowitz, 2008), so only a brief discussion is given here. Based on a limited period of sampling, a relationship is obtained between Q and SSC. The most common form is a log–log power function of the form:

High-resolution monitoring method: interpolation approaches

SSC = aQb

The most effective and accurate method of deriving total SSLs is field intensive and involves automated high-resolution, quasi-continuous sampling or monitoring, at sampling intervals that are short enough to ensure that SSC or flow does not change significantly (i.e. it is “safe” to interpolate Q and SSC between sample values). This interval is often set at between 10 and 30 minutes (commonly 15 minutes) in small- to medium-scale catchments in which flows and suspended sediment fluxes can change quickly (e.g. drainage basin area of < 1000 km2). Examples of such high-resolution studies in the UK include those of Lawler (2005b, 2006) and Old et al. (2003, 2006), and an example of such a study in Iceland is provided by Lawler et al. (2003). The monitoring period should also

(Equation 5.4)

In Equation 5.4, a and b are coefficients determined empirically through regression. The sediment rating relationship is then applied to a high-resolution discharge time series to produce values for the SSLs transported over a given period. Separate sediment rating curves are sometimes used to cater for (1) lower or higher flows; (2) any significant differences between rising and falling limbs of a hydrograph; or (3) seasonal variations in SSC for a given flow (see Figure 5.11). However, the sediment rating curve method has attracted controversy, partly because the relationship between Q and SSC is normally characterised by huge scattering and

47

SILTFLUX Literature Review

uncertainty (see, for example, Figure 5.11), and the SSC values used are, of course, estimates. Ferguson (1987) recommends the use of a statistical correction factor if log–log sediment rating curves are used.

from 0.001 to 5 m3/second per km2, SSCs varied from 1 to 5000 mg/L. The value of exponent b (from Equation 5.4) for British rivers ranged from 0.3 to 2.0, and typically, was approximately 1.2. It was tentatively suggested by Walling and Webb (1981a) that lowland clay catchments have high a values and low b values and, conversely, upland catchments with resistant lithology are characterised by low a values and high b values.

Figure 5.11 (Walling and Webb, 1981b) shows a cluster of log–log regression lines for 17 rivers (many in western or upland Britain), which demonstrates that for discharges normalised for catchment areas ranging

48

6

Suspended Sediment Concentrations, Fluxes and Yields

6.1 Introduction

6.2 Ireland

Material is delivered from catchments to downstream river reaches, floodplains, reservoirs, lakes, and tidal zones and seas in three main forms: suspended sediment, bedload and dissolved load (solutes). Exceptionally high SSCs can result. For example, a concentration of 500,000 mg/L was observed in the Rio Puerco, New Mexico, USA, by Dunne and Leopold (1978) (i.e. the sample was half sediment by weight). High SSCs, well above 2000 mg/L, have also been sampled in glacial rivers (e.g. Lawler et al., 2003) and a Mediterranean river (see Turner et al. 2008), but in humid temperate environments, concentrations tend to be lower, often < 1000 mg/L. A range of SSC values with respect to discharge is shown for several UK rivers in Figures 5.11–5.12.

6.2.1

Existing research on suspended sediment release to Irish waters

Only a limited number of sediment flux studies have been undertaken in Ireland compared with other Eurpoean countries (Vanmaercke et al., 2011), and no comprehensive countrywide Irish study has yet been undertaken, unlike in, for example, the UK (e.g. Cooper et al., 2008; Walling et al., 2008b). The implementation of the EU WFD (2000/60/EC) has been the principal driver for suspended sediment research in Ireland. The integrated approach, which links water quality to catchment management, that is required according to this legislation has resulted in an increase in the number of studies investigating sediment levels in Irish catchments. Relevant studies are shown in Figure 6.1. All studies, apart from the “Eutrophication from Agricultural Sources” and “Pathways” projects, are ongoing.

In terms of total river loads, suspended sediment is often the most important flux, although its importance varies widely across the globe and over time. Typically, in humid temperate environments, the percentages of different sediment outputs are as follows: 50% suspended sediment; 45% solutes; and 5% bedload (Walling and Webb, 1981a). Coarse bed material may move on only a few occasions per annum in such catchments. However, in mountain and, especially, glacial rivers, bedload fluxes can account for 50% of total output. In these types of rivers, high and variable discharges and stream powers, and an almost limitless supply of loose, coarse gravels in the subglacial and extraglacial zones and braid plains, combine to allow the drainage system to access and transport large volumes of bedload material downstream. However, in these cold environments, in which chemical reaction rates are low, and runoff is rapid (reducing solute acquisition opportunities), dissolved loads may be as low as 5% of total sediment output.

Measured Irish sediment yields range from 2.1 to 48.4 t/km2 per year (Table 6.1) and are relatively low in comparison with those reported for other European countries (Vanmaercke et al., 2011). Data related to lake sediments confirm these low sediment levels. As an example, Jordan et al. (2002), in a study of Friary Lough (in Northern Ireland), reported reasonably constant yields of approximately 10 t/km2 per year. Sediment yields throughout the 20th century of between 14 and 58 t/km2 per year were estimated by Huang and O’Connell (2000) in a study of Ballydoo Lough (Connemara). The annual mean SSCs reported as part of the Agricultural Catchments Programme (ACP) (Sherriff et al., 2015) are lower than the annual mean Freshwater Fish Directive guideline value of 25 mg/L. Furthermore, as reported by Sherriff et al. (2015) and Thompson et al. (2014), event-based exceedances in Irish catchments are typically limited to short durations. These, however, can be associated with significant

This chapter summarises some of the available suspended sediment transport data for Irish rivers, and for similar environments in the UK and northern Europe.

49

SILTFLUX Literature Review

Figure 6.1. Catchments that have been studied in recent sediment-related Irish studies. “Eutrophication from Agricultural Sources” refers to EPA project 2000-LS-2.1.1a-M1; “Pathways” refers to EPA project 2007-WQ-CD-1-S1. ACP, Agricultural Catchments Programme; CIT, Cork Institute of Technology; OPW, Office of Public Works data. Table 6.1. Summary of Irish sediment yields reported in the scientific literature Catchment

Area (km2)

Sediment yield (t/km2 per year)

Reference

Corduff

3.34

6.07–22.28

Sherriff et al. (2015)

Mount Stewart

7.52

6.7

Thompson et al. (2014)

Timoleague

7.9

3.95–14.92

Sherriff et al. (2015)

Dunleer

9.4

13.5–41.8

Melland et al. (2012), Sherriff et al. (2015)

Castledockrell

11

2.11–23.1

Melland et al. (2012), Sherriff et al. (2015)

Ballycanew

11.4

6.65–48.39

Sherriff et al. (2015)

Dripsey

15.24

9.8–16.1

Kiely et al. (2007)

Glenamong

17.91

16

May et al. (2005)

Mattock

20.96

44

Thompson et al. (2014)

Clarianna

29.8

8.5

Kiely et al. (2007)

The Oona

84.5

29–41

Kiely et al. (2007)

Owenabue

103

25.6

Harrington and Harrington (2013)

Bandon

424

14.2

Harrington and Harrington (2013)

sediment loadings and can present issues during sensitive periods of potential salmon migration and spawning. These patterns of sediment mobilisation are consistent with those reported by Harrington

and Harrington (2013), obtained from studies of the Bandon and Owenabue Rivers in south-west Ireland. For example, Harrington and Harrington (2010) estimated that 28% of the annual sediment flux of the

50

D. Lawler et al. (2010-W-LS-4)

Owenabue River was transported over only 5% of the time.

Although suspended sediment yields in Ireland are relatively low in comparison with those reported for other European countries (Vanmaercke et al., 2011), Irish catchments are often characterised by the presence of sensitive habitats, including Atlantic salmon spawning grounds and freshwater pearl mussel beds (National Parks and Wildlife Service, 2008), and, therefore, the ecological impacts of elevated sediment levels are particularly significant. Ireland must, therefore, develop adequate regulations for sediment control. Data from Irish sediment research (Figure 6.1) continue to contribute to the characterisation and understanding of the processes that influence the mobilisation and delivery of fine sediments from source to receptor. This knowledge, combined with sediment yield data from Irish catchments, will be essential for the establishment of standards in a regulatory framework for sediment control.

As expected, peak SSCs for agricultural catchments are typically reported during storm events. For example, Harrington and Harrington (2013) reported maximum SSCs of between 837 mg/L and 979 mg/L in their study catchments, which are characterised predominantly by pasture land use. Peak SSCs for the ACP grassland catchments reported by Sherriff et al. (2015) range from 419 to 1020 mg/L. These values are comparable to those reported for catchments primarily used for arable farming, which have a higher potential for soil loss from increased tillage operations that expose bare soils. For example, Sherriff et al. (2015) reported peak SSCs of 221 to 2141 mg/L for this land use type. However, there is an increasing recognition that sedi­ ment loadings from pasture lands can be increased by poor animal husbandry and land management. This includes, but is not limited to, overgrazing and soil poaching by livestock, particularly around gateways and at drinking points on land or adjacent to rivers. These factors have been shown to adversely affect Irish freshwater systems (Conroy et al., 2016). Bank erosion caused by livestock poaching is also problematic in Irish catchments and is one of the factors that contributed to elevated sediment loadings in the Bush catchment, Northern Ireland (Evans et al., 2006).

6.3

Britain and Northern Europe

Relationships between suspended sediment yields, bedload and catchment areas for British rivers were derived by Sear et al. (2003) and are shown in Figure 6.2. SSLs, in t/year, tend to be higher than bedload fluxes, and are linked to the area of the catchment (A), in km2, as shown in Equations 6.1 (for small “sediment source” catchments of < 100 km2) and 6.2 (for large catchments of < 100 km2):

Moreover, the erosion of upland peatland because of overgrazing by sheep can contribute to sediment loadings (Allot et al., 2005; May et al., 2005). This is a particular problem in upland peat catchments in Ireland. Significantly, sediment from these catchments is characterised by high levels of particulate organic carbon (Ryder et al., 2014), which can be more damaging to the ecological status of receiving waters than other types of sediment.

SSL = 11.64A1.16 (n = 60, r2 = 0.63)

(Equation 6.1)

SSL = 31.04A1.04 (n = 44, r2 = 0.48)

(Equation 6.2)

Note that the exponent for large catchments (i.e. 1.04) is lower than that for smaller catchments (i.e. 1.16); this probably reflects the declining erosion (sediment production) and increasing sediment storage opportunities in a downstream direction. Cooper et al. (2008) and Collins et al. (2009) have produced some very useful suspended sediment yield data, which map to over 100 British catchments in south-western England, southern and mid-Wales, the Midlands and the Pennines [including many data from the Natural Environment Research Council’s Land Ocean Interaction Study (NERC LOIS) programme], and southern Scotland. This compilation, based on lake sediment accumulation data and fluvial suspended sediment transport estimates, shows that suspended sediment yields vary from 2 to 160 t/km2

An increase in suspended sediment loadings can also occur during forestry clearfelling (Allott et al., 2005, Rodgers et al., 2011) and windrowing (Clarke et al., 2015) operations. Increases in peak SSC, from 88 mg/L before windrowing to 502 mg/L during windrowing operations, were reported by Clarke et al. (2015) and illustrate that forestry operations can produce SSC levels that are comparable to the upper range of those for agricultural catchments.

51

SILTFLUX Literature Review

1000000

SUSPENDED LOAD

100000

Sediment Load (tonnes/year)

10000

RESERVOIRS

BED LOAD

1000

100

SOIL EROSION

BED LOAD

10

SOIL EROSION SUSPENDED LOAD

1

0 0.001

0.01

0.1

1

10

100 Catchment Area (km2 )

1000

10000

2

From catchments with Area < 100 km 1.08

Bed

L=

5.85 A

Suspended

L = 11.64 A

1.16

(n=33, r

2 2

(n=60, r

= 0.31) = 0.63)

2

From catchments with Area > 100 km Bed Suspended

L=

1.16

2.50 A

1.04

L = 31.04 A

(n=7,

2

r

2

(n=44, r

= 0.41) = 0.48)

Figure 6.2. Observed sediment yield (bedload and suspended load) data as a function of catchment area for UK rivers (after Sear et al., 2003). per year, with the highest yields observed in northern Britain (in the Pennines and the Lake District) and in northern and mid-Wales .

sediment yields and the Walling typology (Walling et al., 2008b): the highest suspended sediment yields were found for upland, low-impact catchments (types 1 and 2; see Table 6.2). For the WFD typology, 16 catchment types (of the 44 types in the WFD definition) were represented in the British dataset. The highest yields tended to be generated from mid-altitude, extrasmall and small catchments (Table 6.3).

Cooper et al. (2008) attempted to link suspended sediment yields to catchment characteristics and three different typologies (the Walling system, the WFD system and a new typology created by the authors of the study) within a combined geographical information system and digital elevation model framework. Table 6.2 shows the relationship between suspended

The new typology developed by Cooper et al. (2008) is shown in Table 6.4; from this table, it is clear that

52

D. Lawler et al. (2010-W-LS-4)

Table 6.2. The Walling catchment typology: links to suspended sediment yield (adapted from Cooper et al., 2008). Type

Altitude

Impact

Size

Yield (t/km2 per year) < 3

3–7

7–23

23–56

56–90

> 90

Total

1

2

2

2

5

9

21

2

3

1

Mid

Low

XS

2

Mid

Low

S

3

Mid

Agric

XS

4

Mid

Agric

S

6

Low

Low

XS

7

Low

Agric

XS

7

3

10

8

Low

Agric

S

3

2

6

7

1

19

9

Low

Agric

M

3

4

9

7

2

10

Low

Agric

L

1

4

1

3

11

Low

Urban

XS

1

12

Low

Urban

S

1

1

1

1 1

3

4

3

3

1

1

2

27 9 1 3

Impact is characterised as “low”, “agricultural” or “urban”. The replacement of “geology” by “impact” in the typology of Walling broadly replaces “siliceous” and “organic” with “low impact”, and “calcareous” with “agricultural”, with the “urban” classification being new. The catchments are classified according to mean altitude (low: < 200 m; mid: 200–600 m; high: > 600 m) and size (XS: < 10 km2; S: 10–100 km2; M: 100–1000 km2; L: 1000–10,000 km2; XL: > 10,000 km2).

Table 6.3. The WFD catchment typology: links to suspended sediment yield (UKTAG, 2008). This typology is based on three catchment characteristics (altitude, size and geometry) (adapted from Cooper et al., 2008) Type

Altitude

Size

Geology

1

Low

S

SI

2

Low

S

CA

4

Low

M

SI

5

Low

M

CA

8

Low

L

CA

10

Mid

S

SI

11

Mid

S

CA

13

Mid

M

SI

14

Mid

M

CA

16

Mid

L

SI

17

Mid

L

CA

37

Low

XS

SI

38

Mid

XS

SI

40

Low

XS

CA

41

Mid

XS

CA

44

Mid

XS

OR

Yield (t/km2 per year) < 3

3–7

3

2

3

4 1

Total 7–23

23–56

56–90

2

1

1

4

8

1

1

3

2

4

1

4

1

3 1

1

2

1

2

2

5 3 7

1

3

2

3

1

1

2 1

13 6

2

1

1

4 17

3 1

> 90

2

1

4

2

3

2

3

8

2

2

4

1

1

1

6

3

16

6

The catchments are classified according to mean altitude (low: < 200 m; mid: 200–600 m; high: > 600 m), size (XS: < 10 km2; S: 10–100 km2; M: 100–1000 km2; L: 1000–10,000 km2; XL: > 10,000 km2) and dominant geology. Geology is classified by the British Geological Survey as SI, CA, OR or SA (Kinniburgh and Newell, 2003). The SI and CA classes relate to solid geology, while the OR and SA classes are based on near-surface characteristics. CA, calcareous; OR, organic; SA, saline; SI, siliceous.

53

SILTFLUX Literature Review

Table 6.4. The new “Natural England” typology: links to catchment suspended sediment yield (adapted from Cooper et al., 2008) Altitude

Permeable/ impermeable

Geology

Index

Low

Permeable

Chalk

LPC

Low

Permeable

Other

LPO

Low

Impermeable

Other

LIO

Low

Impermeable

Peat

LIP

High

Permeable

Other

HPO

High

Impermeable

Other

HIO

High

Impermeable

Peat

HIP

Yield (t/km2 per year) < 3

3–7

7–23

6

7

3

1

16 3

Total 23–56

56–90

> 90

19

5

2

5

8

16

1 1

1

17

1

2

1 2

1

2

1

4

43

2 5 2

9

16

The first letter in the index classification refers to altitude with H and L referring to high (> 330 m) and low (< 330 m) altitudes, respectively. The middle letter relates to the Standard Percentage Runoff with P and I referring to permeable and impermeable catchments, respectively. The third letter refers to the Hydrology of Soil Types soil classification, with C, P and O describing chalk, peat and other soil classes, respectively. HIP, high impermeable peat; LIP, low impermeable peat.

of western, northern and central Europe: these catchments accounted for half of the low sediment yields (< 40 t/km2 per year) in the database, and approximately 80% of the sediment yields of < 200 t/ km2 per year.

the two highest classes of suspended sediment yields are delivered from catchments classed as “hip” (high impermeable peat; 16 catchments) and “lip” (low impermeable peat; two catchments). The authors also use this new typology to estimate suspended sediment yields across England and Wales from catchment typology information generated by national databases of catchment characteristics on a 1-km2 grid scale (Figure 6.3).

Vanmaercke et al. (2011) also found a statistically significant (p < 0.05) inverse relationship (as others have in smaller-scale studies) between catchment area and sediment yield. This relationship emerged for most climatic zones, including zones classified as “Atlantic”, according to their classification, which are of most relevance to Irish environments. This suggests that sediment storage opportunities increase downstream, as steep topography gives way to relatively low and flat landscapes where sediments can be deposited at slope bases, in floodplains or on channel bars and beds. Generally, they found stronger relationships between climate and sediment yield, than between topographic indices and sediment yield.

However, British suspended sediment yields are low by world standards and compared with mountain and glacial catchments in northern Europe. For example, Lawler (2003) showed that the approximately 100km2 glacierised basin of Jökulsá á Sólheimasandi in southern Iceland delivers suspended sediment yields of 10,000 t/km2 per year, and suspended sediment yields for glacial catchments of similar orders of magnitude have been analysed by Gurnell et al. (1996). However, the relatively low sediment fluxes of well-vegetated humid temperate environments can still lead to significant negative impacts on fluvial hydroecology (see Chapter 4).

In a survey of downstream water clarity in three unregulated catchments in New Zealand and Wisconsin, USA, Julian et al. (2008a) found that turbidity increased (and clarity decreased) along the channel length to a zone approximately 70% along the continuum; after this point (i.e. further downstream), turbidity declined slightly (Figure 6.3). This is consistent with the mid-basin peak in-stream power modelled by Lawler (1992) and validated by Barker et al. (2009), thought to result from enhanced bed instability and river bank erosion.

An excellent compilation and analysis of sediment yield data for the whole of Europe, including Ireland, has recently been published by Vanmaercke et al. (2011). This was based on a dataset of over 29,000 catchment-years and 1794 sites (507 reservoirs and 1287 gauging stations), for drainage basin areas ranging from 0.01 to 1,360,000 km2. They found that the lowest sediment yields were in the flat areas

54

Bean attenuation coefficient, c (m–1)

D. Lawler et al. (2010-W-LS-4)

Normalised river length (% of total length) Figure 6.3. Downstream changes in optical water quality in six rivers in Wisconsin (USA) and New Zealand (Julian et al., 2008a).

55

7

Storm-Event and Seasonal Suspended Sediment Dynamics

7.1

Storm-Event Suspended Sediment Dynamics

of the flow maximum, before a quick decline. This is the classic response described by the “first-flush model” (Figure 7.1 and Figure 7.2). If SSC is plotted against synchronous discharge, a classic positive (i.e. clockwise) hysteresis relationship often emerges, as is shown by, for example, Walling and Webb (1981a) and Walling (1997).

Suspended sediment yields are useful, but unless attempts are made to understand the sources, delivery, processes and dynamics of sediment transfer from sources to rivers, then management initiatives aimed at sediment control could be less targeted; it may be more difficult to engage potentially sceptical stakeholders and farmers, and such initiatives may be problematic to implement and possibly fund. A consideration of sediment dynamics is thus crucial for the understanding of the sediment system, and thereby informing any management solutions. This section summarises the typical dynamics of sediment responses in river systems.

7.2.1

Intra-storm hysteresis effects

A common attribute of catchment sediment systems is positive hysteresis, by which the maximum SSC precedes the flow peak (e.g. Stubblefield et al., 2007). Thus, SSCs are already falling by the time of the flow peak, suggesting that channel hydraulics are not the only variables that influence suspended sediment transport.

Although significant suspended sediment pulsing can occur independently of Q variation in some catchments (e.g. glacierised basins; see Lawler and Brown, 1992; Lawler, 2005b), numerous empirical studies have shown that, in humid temperate environments, most suspended sediment is transported during storm events (e.g. Walling, 1974; Walling and Webb 1981b; Old et al., 2003, 2006; Owens, 2009). Indeed, in many catchments, most suspended sediment flux is transported in just a few events, mainly during and soon after rainstorms. For example, Old et al. (2003) showed that 40% of annual SSL in the Bradford Beck urban stream is transported during only approximately 1% of the year, and Horowitz (2008) reported that > 94% of the fluvial fine sediment transported in Atlanta, USA, occurs during storm events occupying < 20% of the time. It is, therefore, vital that, in any suspended sediment investigation, maximum resources are devoted to high temporal resolution monitoring and/or sampling of SSCs during such highflow (or other turbidity-generating) events.

Walling (1974), and many others since, have argued that this positive hysteresis is driven by the early flushing of readily accessible, but finite, supplies of loose erodible fine sediment in the catchment. This applies particularly to supplies from river bank erosion, which can be more pronounced early in the storm, when rising flood waters “attack” bank materials that have previously been weakened by weathering and preconditioning processes, such as freeze–thaw action (Hill, 1973, study in Northern Ireland; Lawler et al., 1997; Lawler, 2005) (see Chapter 2). Loose, highly erodible, exposed sediment sources are often widely available on catchment surfaces (e.g. from arable fields and poaching sites), and these supplies can be relatively quickly entrained during the start-up phase of the storm. However, these supplies are also quickly depleted – a concept known as “sediment exhaustion” – leaving little available fine sediment for transport in the later stages of the hydrograph. Positive hysteresis may also arise because of “baseflow dilution”, that is, waters arriving later in the storm via sub-surface baseflow routes can be sediment poor because the potential for erosion and entrainment of fine sediment is limited by low throughflow velocities.

7.2 Hysteresis One of the most common sediment-related responses to occur during storm events, especially in small catchments, is the rapid rise in SSCs during the rising limb of the hydrograph, then a peak ahead

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D. Lawler et al. (2010-W-LS-4)

180

100

160

SSC

140

90

Discharge

80

70

Discharge (m3)

120

Discharge

100

60

50 80 40 60 30 40

20

20

0

10

0

5

10

15

20

25

30

Suspended sediment concentration (mg/L)

SSC

0

Time (hours)

Suspended sediment concentrations (mg/L)

180

160

suspended sediment leads flow

140

120

100

80

60

sediment exhaustion

40

20

0

0

10

20

30

40

50

60

70

80

90

100

Discharge (m3/s)

Figure 7.1. SSC response dynamics: SSC leading the flow (top panel) and the classic positive hysteresis and first-flush model of sediment dynamics (bottom panel) (after Lawler et al., 2006).

7.2.2

This first-flush concept is now enshrined in numerical sediment transport models and those for other contaminants [e.g. the Storm Water Management Model of the United States Environmental Protection Agency (USEPA); see Akan and Houhhtalen, 2003]. Such models allow for the accumulation of sediments and/or contaminants between storm events; these supplies are then very quickly entrained by rainfall or river flows early in the storm event, leaving few supplies for the recessional limb of the hydrograph.

Inter-storm hysteresis effects

Positive hysteresis is also evident on weekly timescales if multiple storm events occur. For example, Figure 7.2 shows that for a sequence of storm discharge rises, successive SSC peaks tend to progressively decline, despite flow maxima that are of similar magnitudes. This can also lead to multiple clockwise hysteresis loops.

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SILTFLUX Literature Review

3

Discharge (m )

2000

20

Discharge 1000

10

suspended sediment concentration

0

after Walling et al 1981

27

26

25

28

Suspended sediment conc. (mg/L)

3000

30

December 1974 Figure 7.2. Classic suspended sediment dynamics in response to storm-event discharge changes on the River Dart, south-west England (after Walling and Webb, 1981a).

7.2.3

Negative, anticlockwise hysteresis

hydrograph limbs is largely restricted by highly impervious urban surfaces, such as concrete, glass, steel, tarmac, brick and stone.

It is important to stress that the first-flush model, although commonly applicable, is not entirely ubiquitous. For example, studies by Heidel (1956) and Asselman et al. (2003) on the Rhine, and Lawler et al. (2006) on the urbanised River Tame, have shown that negative, anticlockwise hysteresis, in which peak SSCs occur just after the flow maximum (e.g. Figure 7.3), can also be a typical response. For the Tame, although some clockwise hysteresis responses emerged, anticlockwise hysteresis was by far the most common response (Figure 7.4). Lawler et al. (2006) quantified the magnitude and direction of hysteresis for these events with a simple hysteresis index, and suggested a range of hypotheses to account for the anticlockwise hysteretic behaviour, including:

Thus, hysteresis, both positive and negative, is very common in fluvial sediment transport systems. One implication of hysteresis is that it makes predictions of SSCs from Q using a rating curve approach very difficult (see Section 5.9).

7.3

Seasonal Changes in Suspended Sediment Fluxes

For humid temperate environments, there have been surprisingly few studies devoted to seasonal changes in SSCs or fluxes. This is in contrast to the numerous investigations of seasonal changes in glacierised catchments, in which the high flows of the summermelt season are matched by very high SSCs and loads (e.g. Lawler et al., 2003; Old et al., 2005; Stott et al., 2009).

1. sediment-wave versus water-wave translation differences (Heidel, 1956; Marcus, 1989; Bull, 1997; Knight, 2005), i.e. wave celerity phenomena;

Exceptions include the work of the Exeter group led by Walling (Walling, 1974, 1988; Walling and Webb, 1981a) on Devon rivers. These studies showed that, perhaps surprisingly, summer SSC values tend to be higher than winter values (e.g. Figure 5.11). This was thought to be because the inter-arrival time for summer storms is usually much longer than that for winter storms, allowing a longer period for the preparation of fine sediment supplies ready for removal from catchment and river bank surfaces and delivery to the river mainstem. Such preparation processes may

2. distal sediment sources that take a long time to arrive in the channel, as Heidel (1956) observed; 3. biofilm break-up late in the hydrograph, releasing trapped sediment; 4. combined sewer overflows, which spill late in storm events. One possibility is the lack of baseflow dilution in urban catchments, because the necessary basin water storage for strong baseflows to occur on recessional

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D. Lawler et al. (2010-W-LS-4)

River Tame at Water Orton

50

500

45 40

400

Discharge (m3)

30

300

25 200

20 15 10

Turbidity (FTU)

35

100

5 0

0

12

36

24

0

Time (hours) Figure 7.3. The typical suspended sediment dynamic response in the urbanised River Tame catchment (Birmingham, UK) is negative, anticlockwise hysteresis, in which peak SSCs occur just after the flow maximum (after Lawler et al., 2006).

500 3 April

Turbidity (FTU)

400 4 April 300

200

100

0

0

1

2

3

4

5

6

Discharge (m /s) 3

Figure 7.4. Clockwise hysteresis and anticlockwise hysteresis (the most common loops in the Q– turbidity relationship) for the River Tame, Birmingham (after Lawler et al., 2006).

include desiccation-related cracking of field soil and, especially, bank materials (see, for example, Lawler et al., 1997). This may be exacerbated by a tendency for intense convective storms and storms with greater raindrop diameters to strike in summer. Seasonal suspended sediment hysteresis was also noted for mountain streams that drain to Lake Tahoe (California– Nevada border, USA) by Stubblefield et al. (2007).

Although winter suspended sediment concentrations may be lower, suspended sediment yields may be higher, given the higher river discharges and the tendency for rainstorms to fall on bare arable fields from which soil particulates can be entrained more easily. Kemp et al. (2011) show a useful example of this for agricultural catchments, and with respect to seasonally changing cultivation patterns and risk periods for fish at different life stages.

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SILTFLUX Literature Review

7.4

Longer Term Changes

Lawler et al. (2003) found that sediment yields in three Icelandic glacierised basins were amongst the highest in the world, at approximately 10,000 t/ km2 per year. However, in general, these yields decreased during the 1973–1992 period examined, and, although meltwater season sediment transport increased, this appeared to be more than offset by decreased river flows and sediment transport in spring and autumn, leading to a net reduction in suspended sediment yields. Walling (2005) identified a trend for decreasing suspended sediment yields for world river basins, which was ascribed to the impact of dams and reservoirs on sediment storage. Horowitz (2010) documented a decline in suspended sediment fluxes in the Mississippi River from 1981 to 2007, and suggested that this resulted from a reduction in sediment supply (rather than discharge) associated with the emplacement of artificial structures, but also improved land management. Conversely, fine sediment generation appears to be increasing in some catchments, including that of Lake Tahoe (Jassby et al., 1999).

Changes in suspended sediment yields over decadal and century timescales have been considered, but less intensively. For example, Walling et al. (2003) studied sediment cores from lakes and reservoirs in order to reconstruct suspended sediment yields over the previous 150 years for the Ouse and Tweed river catchments in northern England. The sediment yields for the Ouse generally increased after 1963, which reflects the intensification of agriculture at this time. For the Tweed, suspended sediment yields were higher in the earlier period of the study, which the authors ascribed to afforestation and switches from pasture to arable land use in the 1940s and 1950s. Climate apparently had little effect on either catchment. Walling and Fang (2003) analysed data from a selection of the world’s rivers and reported a decreasing trend for sediment concentrations; they concluded that reservoir sedimentation played the largest role in controlling sediment flux to oceans.

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8

Effects of Land Use and Climate Change on Sediment Fluxes

8.1 Introduction

8.2

The sediment yield of a basin depends on numerous factors including rainfall, storm runoff, basin relief, size and morphology, tectonics, bedrock lithology and structure, topography, soil type, land cover/use and climate (Milliman and Syvitski, 1992; Gordeev et al., 1996; Inman and Jenkins, 1999; Farnsworth and Milliman, 2003; Molina et al., 2008). This chapter looks at the impact of climate change and land use on sediment fluxes. Land use and climate change (along with runoff) are the only factors that influence sediment yield that are subject to change over relatively short periods. It is therefore important to understand and be able to predict the impacts of these factors on sediment yields.

Climate Change with Particular Reference to Ireland

Over the course of the 20th century, the globalaverage surface temperature increased by 0.6°C (±0.2°C) (Betts, 2002; McElwain and Sweeney, 2003) and overland precipitation increased by 1% (Kiely, 1999). Global mean temperatures are predicted to have increased by between 1.4°C and 5.8°C by the end of the 21st century, with a corresponding increase in global mean precipitation of 2.4% per 1°C increase in temperature (Betts, 2002; Jasper et al., 2004; Fealy and Bates, 2009). Enhanced hydrological cycles (leading to increases in the rates of precipitation and the rates of evaporation and runoff) are expected for parts of the northern latitudes, including western Europe (Kiely, 1999). In recent decades, there has been an increase in mean precipitation, as well as more frequent heavy precipitation events over the majority of land in northern Europe (Steele-Dunne et al., 2008).

Every catchment’s sediment yield will change uniquely in response to alterations in climate and land use. This often leads to difficulties with regard to making general statements on how a particular change in a certain characteristic of a catchment will affect that catchment’s sediment yield. As noted above, land use and climate change are only two of a number of variables that influence the sediment yield of a basin. There are numerous examples in the literature of increasing and decreasing sediment yields of different catchments undergoing the same change in land use (e.g. Bakker et al., 2008; Hunter and Walton, 2008). Some basins are better able to cope with changes by remobilising stored sediment in times of deficit (Walling, 1999), while some intensive land use changes may occur in areas of poor connectivity with the river channel, and therefore will not affect the sediment yield of the basin at all. It is important to keep this in mind when trying to identify the likely effects of land use and climate changes on sediment yield.

While it is uncertain whether or not the trend in Irish temperatures over the previous century matches those identified by the Intergovernmental Panel on Climate Change, the Irish climate, in general, is following similar projections to those predicted by global climate models (McElwain and Sweeney, 2003; Fealy and Bates, 2009). Ireland and Great Britain have experienced an enhancement of hydrological cycles and a stronger seasonality than before, in precipitation and streamflow, is becoming evident (Kiely, 1999; Betts, 2002); furthermore, even more apparent seasonality is predicted for the 21st century (Arnell and Reynard, 1996; Pilling and Jones, 2002; Fealy and Bates, 2009). February, March and October are the wettest months and appear to be getting wetter, while May, August and September tend to be the driest (McElwain and Sweeney, 2003), although extreme events of higher magnitudes are occurring more frequently

To accurately determine why sediment fluxes change, it is necessary to have a detailed understanding of past and present information on sediment fluxes, land uses and climatic conditions (Houben et al., 2006).

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SILTFLUX Literature Review

in September and October (Kiely, 1999). Winter precipitation in Ireland is expected to increase by between 6% and 13% over the next 40 years, and summer precipitation is expected to remain unchanged or decrease by up to 8% (Betts, 2002).

Thodsen et al. (2008) discussed the secondary effects that climate will have on sediment yields. An increase in temperature will change the growing and sowing seasons for many crops, in turn changing the temporal and total erodibility of the soil in which they grow. Growing periods for river macrophytes and plants will be extended and this will alter the sediment transport dynamics within the river channel.

There is some debate regarding the possible future spatial variations in precipitation in Ireland (Fealy and Bates, 2009), but Kiely (1999) noted that there was a significant increase in precipitation on the west coast of Ireland since 1975, with little or no increase in the east. Charlton et al. (2006) predicted a decrease in runoff in the east and south-east, and stated that an increase in the north-west is likely.

The majority of stream flow/runoff values, and hence sediment yields, predicted for the 21st century were based on the assumption that catchment characteristics will remain constant. This assumption is unrealistic as land uses may, and inevitably will, change in the future; such changes will alter the runoff behaviour in catchments and the associated sediment delivery to water bodies.

With regard to the possible future combination of temporal and spatial variations in Irish precipitation, McElwain and Sweeney (2003) and Charlton et al. (2006) have acknowledged that there are likely to be increases in the north-west in winter and decreases in the south-east in summer.

8.3

Moderate land use can double or triple the sediment yields of a basin, while intensive land use can increase yields by an order of magnitude (Saunders and Young, 1983).

Observational records suggest that both the frequency of occurrence and the intensity of extreme events is increasing (Fealy and Bates, 2009). Kiely (1999) provided evidence that the number of extreme events in Ireland has increased since 1975 (although this probably has more to do with the natural climate variability/influence of the North Atlantic Oscillation rather than global climate change). Murphy and Charlton (2008) and Fealy and Murphy (2009) predicted that the frequency of extreme events (of both high and low magnitudes) will increase over the course of the 21st century.

8.2.1

Land Use

Erosion processes create the supply of sediment, while factors and processes such as basin slope, runoff, and hillslope/river channel linkage (Foulds and Macklin, 2006) influence the transport of this sediment. Agricultural land use changes, changes in forest management practices and climate change have led to an increase in erosion rates (Scheurer et al., 2009). Global soil erosion potential is estimated to have increased by 17% during the 20th century because of the development of arable land (Yang et al., 2003).

Effects of climate change on sediment fluxes

The amount of sediment that actually reaches the river channel, relative to the amount available for transport, is denoted by the SDR (Walling, 1997; Asselman et al., 2003); this value is unique for each catchment. Some land conditions are more susceptible to erosion than others, while some are better than others at transporting sediment. Changes in land use/cover alter the erosion and transport processes of a catchment; this, in turn, will alter the SDR and, ultimately, the sediment yield. Table 8.1 shows some of the different effects that land cover changes can have on the hydrological characteristics of a catchment.

More intense rainfall patterns are expected to result in increased runoff, which, in turn, will transport more non-point source runoff into waterways (Kiely, 1999; Dornblaser and Striegl, 2009) resulting in an altered nutrient balance and greater levels of turbidity (Penck et al., 2009). More importantly, most sediment is delivered to water bodies during low-frequency storm events; therefore, an increase in the frequency of events is likely to lead to an increase in sediment concentrations/loads (Hunter and Walton, 2008; Scheurer et al., 2009).

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Table 8.1. Some effects of land cover changes on catchment characteristics (modified from Bronstert et al., 1999) Change of land cover

Affected catchment characteristics

Urbanisation (increase of impermeable areas)

No soil storage Accelerated runoff concentration

Deforestation

Less canopy storage and root depth Less litter storage Reduced infiltration capacity

Forest damage (by acid rain)

Less canopy storage

Consolidation of farmland (re-allocation and rationalisation)

Less soil and canopy storage Reduced surface storage Accelerated runoff concentration

Intensive agricultural usage (soil compaction, use of heavy machinery)

Reduced infiltration capacity Reduced ponding storage Accelerated runoff concentration

Change in vegetation composition (triggered by climate change or by anthropogenic activities)

8.3.1

Changed canopy storage and root depth Changed soil parameters, e.g. infiltration capacity

Dominant global agricultural land uses/ covers and influences

et al., 2007). Bakker et al. (2008) reported substantial decreases in sediment export to rivers in Europe because of the conversion of marginal agricultural arable land to forests, grasslands and scrublands.

Agriculture can have a major impact on sediment fluxes in a basin. The clearance of protective vegetative land cover, for example by the use of machinery on fields (Scheurer et al., 2009) or the degradation of land by animal pastures, leaves the soil exposed to erosion and creates drainage pathways over the land, allowing for more rapid and larger volumes of sediment transport. Agricultural practices can affect a landscape’s water retention potential and its infiltration capacity (Bronstert et al., 1999).

8.3.2 Urbanisation Urbanisation within a basin will induce a hydrological change, as natural land is replaced by impermeable or hard cover layers; this will increase runoff and alter the flood peak of the basin. These changes have implications for the sediment yield of the basin: soil previously available for erosion may be concealed, thus reducing erosion processes and leaving less sediment available for transport.

Ward et al. (2009) found that in the Meuse catchment (Western Europe), the primary factor controlling longterm changes in sediment yield was the conversion of forests to arable land. Furthermore, the expansion and intensification of agriculture, which consumed and decreased basin buffer strips, is thought to have increased fine sediment loads in rivers in Alpine regions (Scheurer et al., 2009).

A study carried out by Inman and Jenkins (1999) identified that the extensive hard cover of the streets and river channels of the Los Angeles urban area has significantly reduced the sediment yield. Urbanisation of a basin in California led to a 20-fold decrease in the SSC, with respect to discharge, in the river by a non-linear dilution process due to changes in the hydrological regime (Warrick and Rubin, 2007).

It is assumed that changes in agricultural management and precipitation patterns have increased the sediment yield in the Danube catchment by 30–50% (Scheurer et al., 2009), while the sediment dynamics of the Waipaoa River in New Zealand were severely affected by the large-scale clearance of forest and the subsequent conversion of the land to pasture during the late 19th and early 20th centuries (Gomez

However, large increases in sediment yields and concentrations can occur in basins undergoing urbanisation, as a result of construction activity (Chen, 1974; Walling, 1974; Warnock and Lagoke, 1974). Chen (1974) found that erosion rates were over 50-times higher in urbanising basins compared

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with natural basins in the eastern USA. SSLs in Canadian streams were three- to five-times higher in basins undergoing urbanisation than in rural basins in neighbouring streams of approximately equal size (Warnock and Lagoke, 1974). Walling (1974) found that sediment loads and concentrations in a basin near Exeter, UK, increased by 5- to 10-fold and five-fold, respectively, during construction in an area comprising 25% of the basin.

erosion, landslides and mudslides, which can cause further elevations of SSCs by, in some cases, up to nine times that of the undisturbed levels (Fredriksen, 1970). The sediment yield rates of a basin in New Zealand, which was tracked, harvested and burnt, were eightfold higher than the yield rate of a nearby forested control basin (O’Loughlin et al., 1980). The first flush of sediment from a forested basin that had undergone patch-cutting and road construction had a peak SSC that was 250-fold higher than would have been expected from the basin if it had remained undisturbed, and sediment concentrations consistently remained above pre-deforestation levels (Fredriksen, 1970). In Ireland, conifer tree harvesting and windrowing, in preparation for replanting, resulted in elevated episodic inputs of sediment to watercourses that exceeded water quality standards, with the largest releases near the end of the operations (Clarke et al., 2015; Kelly-Quinn et al., 2016).

8.3.3 Forestry Forested catchments have the best natural water storage capacity and lowest runoff yields of all catchment types (Bronstert et al., 1999). Sediment is often “trapped” (deposited) in forests, breaking the link between flow paths and the river channel (Van Rompaey et al., 2002). The unique features of forest cover, such as root systems, litter layers and canopies, are much more effective than most agricultural crops at protecting soil from water erosion (Chang et al., 1982).

Aside from the direct effects of deforestation, it has been shown that the release of suspended sediment into streams because of secondary sources, such as forest roads, tracks and landings, resulted in the most widespread water quality problems in basins in the USA in circa 1970 (USEPA, 1975; Anderson et al., 1976; Swanston and Swanson, 1976). It should be noted, however, that Lees et al. (1997) found that sediment yields can increase significantly immediately after afforestation and claim that other UK studies have shown order-of-magnitude increases in sediment yield relative to pre-afforestation levels; however, these instances are rare and probably involve a spatial element.

Piégay et al. (2004) noted that afforestation appears to reduce sediment supply. During a study in France, Liébault et al. (2005) found that an unvegetated basin frequently had SSCs higher than 300 g/L, whereas a forested basin of the same size had a maximum SSC of 35 g/L. The sediment yield of a catchment can increase significantly (by up to 1300%) with increasing rainfall after deforestation (Foulds and Macklin, 2006). As with urbanisation, deforestation results in a change in the hydrologic response of a catchment (Bronstert et al., 1999). Deforestation can leave land susceptible to

Table 8.2. Total net rainfall, runoff and soil loss resulting from 30 storms between 28 May 1980 and 27 February 1981 in Nacogdoches, Texas (modified from Chang et al., 1982) Forest condition

Net rainfall (mm)

Runoff (mm)

Soil loss (kg/ha) SUS

DEP

Total

Undisturbed forest

359.6

7.4

4.2

6.5

10.7

Thinned forest

377.8

17.0

11.6

5.5

17.1

Clearcut (a)

430.4

41.7

42.5

113.3

155.8

Clearcut chopped

459.4

81.4

75.8

189.2

265.0

Clearcut KG bladed

459.4

137.3

2201.6

1260.0

3461.6

Clearcut cultivated

459.4

119.0

1009.4

2414.0

3423.4

DEP, deposited sediment; SUS, suspended sediment.

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Table 8.2 shows the effects of various destructive forest practices on runoff and soil loss. These data prove that the sediment supply and transport levels increase with an increasing intensity of deforestation.

8.3.4

plot in Germany. However, strong gully formation can occur in unmanaged grasslands (Bakker et al., 2008), which are more sparsely covered than managed grasslands. The widespread conversion of arable land to shrubland in Amendoeira, Portugal, during the 1960s resulted in large decreases in sediment export (Bakker et al., 2008). In a study by Molina et al. (2008), it was found that surface vegetation exerted a first-order control on sediment yield. O’Loughlin et al. (1980) discussed the benefits of a riparian strip with regard to reducing sediment export to river channels.

Other land uses and issues: fire, mining and population changes

The land uses discussed so far are the most common and general in global terms. However, there are many other events and types of land use that can change the land. For example, fire, grassland, shrubland, mining and changes in population can dramatically alter the sediment yield of a basin.

Piégay et al. (2004) and Walling (2006) referred to the negative effects that mining can have on sediment yields in a basin, the latter acknowledging it as a key driver of increased sediment loads. Lewin and Macklin (1987) coined the phrase “active transformation” to reflect the morphological impacts of large-scale sediment releases as a result of mining, although this primarily relates to coarse (rather than fine) sediment loads.

An increase in the frequency of fires will leave more forest soil susceptible to erosion (Dornblaser and Striegl, 2009). Soil exposed by fire can become a source of sediment, particularly coarse sediment (Fredriksen, 1970), as the filtering/trapping action of the original vegetation is negated. Warrick and Rubin (2007) noted that sediment discharge and SSCs increased as a result of an increase in the frequency of upland wildfires in California, while Peart et al. (2009) showed that hill fires increase storm-period SSCs.

Jinze (1991) and Walling (1997) noted the historical impacts of rising populations on sediment loads. For example, Jinze (1991) showed that population increases have, historically, been related to sediment loads, but that under managed conditions this is not the case.

Fiener and Auerswald (2006) highlighted the potential of grassed waterways with regard to reducing sediment delivery: grassed waterways led to a 93% reduction in sediment delivery from an experimental

65

9

Management Implications

Management measures should seek to achieve some combination of (1) a reduction in the sediment load that enters or is carried by a river; and (2) a reduction of the impact of that sediment. It is useful to classify management measures according to the physical process they seek to address. The main physical processes that contribute to SSCs are detachment, mobilisation, advection/dispersion and deposition. Management measures that influence these processes will affect the amount of sediment in rivers.

9.1

Reducing Sediment Load

9.1.1

Rural areas

(b) Erosion by flowing water can be reduced in a field by regrading slopes by, for example, contour terracing, in order to reduce flow velocities. Bank erosion in channels has been studied by Odgaard, (1984), Green et al. (1999), Riedel et al. (2005) and Kessler et al. (2012); in the study by Kessler et al. (2012), lidar techniques were used. Erosion can be influenced by sub-surface flows on hillslopes (Fox and Wilson, 2010) and piping (Hagerty and Spoor, 1989). Van Eps et al. (2004) developed a bank erosion hazard index based mainly on bank angle, root depth and bank material. However, Segura and Booth (2010) also demonstrated relationships between erosion and the geomorphic setting and degree of urbanisation. Bank erosion can be reduced by restoring the natural character of the stream (Chen et al., 2005); by installing special guide vanes (Bhuiyan et al., 2010); by bank-side vegetation (see, for example, Abam, 1993), particularly on bends (Hagerty and Spoor, 1989); by rock armouring (Bogen and Bösnes, 2004); and by using flow redirection structures (Yescas et al., 2011) and structures made from large woody debris (Shields Jr et al., 2004).

Management measures seek to (1) reduce the amount of sediment detached from parent material; (2) reduce the mobilisation of such sediment; and (3) prevent the transport of any mobilised sediment into a watercourse, as outlined below: 1. Sediment is detached from its parent material by any combination of (1) the impact of falling raindrops, (2) erosion by flowing water (Ballantine et al., 2009) and (3) bank collapse: (a) Raindrop impact has been studied by Sharma et al. (1993), Salles et al. (2000), Thompson et al. (2001), Gao et al. (2005) and Kinnell (2011). It may be influenced by wind speed and direction (Erpul et al., 2004). The effects of raindrop impact are reduced by cover crops or mulching (Meyer et al., 1970; Jennings and Jarrett, 1984; Albaladejo Montoro et al., 2000; Grismer, 2007). Lal (1984) has reviewed mulching in the tropics. A special case is the use of greenwaste mulch to control the amount of sediment in surface runoff from landfill sites, although it increased dissolved oxygen concentration (Brodie and Misra, 2009). The effects of raindrop impact are reduced if infiltration occurs (Walker et al., 2007).

2. Mobilisation normally occurs as a result of turbulent flowing water; the turbulence is related to the depth and velocity of flow. Sediment mobilisation is often associated with the mobilisation of phosphorus (Ballantine et al., 2009). The process was modelled (using the PSYCHIC model) by Davison et al. (2008) and evaluated by Stromqvist et al. (2008). A special case of mobilisation is the disturbance of settled sediments; this can have chemical as well as physical effects (Eggleton and Thomas, 2004). 3. The sediment in flowing water can be prevented from reaching a stream by (1) disconnecting the drainage from the stream, (2) retarding the flow so

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9.1.2

that sediment settles or (3) storing the water for a certain period, allowing the sediment to settle. Riparian buffer strips are a common measure used to prevent sediment in overland flow from reaching a stream channel. Wenger (1999) recommended a minimum width of 30 m for these buffer strips, mainly to ensure a healthy biota, but acknowledged that wider buffer strips may be necessary on steeper slopes. Detention pond removal efficiencies are in the range of 50–90% (Hartigan, 1988; Schueler et al., 1992).

Urban areas

Some common measures for reducing sediment export from urban areas to watercourses are listed in Table 9.4. Street sweeping is performed in most urban areas and, depending on the method used, can be effective at removing particles as small as 10 µm in diameter (CASQA, 2003). With regard to efficiency, street sweeping measures can remove up to 70% of total solids (Schilling, 2005). Hallock (2007) reviewed performance data from a number of studies and summarised the sediment removal efficiencies of some commonly adopted measures (Table 9.5). However, quantifying sediment loads can be difficult because of the possibility of deposition and re-suspension of sediment in pipe systems. Rabinovich and Kalman (2011) reviewed the factors that can influence particle transport in pipes.

Measures for reducing the mobilisation of sediments from agriculture and the delivery of sediments to watercourses are summarised in Table 9.1 and Table 9.2. The performances, in terms of sediment reduction efficiency, of selected measures are shown in Table 9.3.

Table 9.1. Reducing mobilisation of sediment from agricultural activities Measure

Reference

Cover crops

Krutz et al. (2009)

Crop residue management or straw mulching

Ritter and Shirmohammadi (2001)

Straw bales or wattles, which can also be used as sediment traps

Baxter (2008)

Contour terracing

Al Ali et al. (2008)

Careful design of farm and forest roads including choice of location far from riparian zone

Anderson et al. (2011)

Table 9.2. Reducing delivery of mobilised sediment to watercourse Measure

Reference

Buffer strip

Dillaha and Inamdar (1997), Wenger (1999), Owens et al. (2007), Grace III and Davis (2010)

Sediment trap (in channel)

Konwinski (1978)

Wetland (natural or constructed)

Zierholz et al. (2001)

Vegetated filter strip

Abu-Zreig et al. (2004), Verstraeten et al. (2006)

Check dams

Shieh et al. (2007), Boix-Fayos et al. (2008)

Silt fences

Barrett and Malina (2004)

Streambank protection

Van Eps et al. (2004), Simon et al. (2008)

Table 9.3. Some estimated sediment reduction efficiencies (Evans and Corradini, 2001) Measure

Estimated efficiency (%)

Crop residue management

64

Vegetated buffer strips

58

Cover crops

15

Terraces and diversions

71

Streambank protection

76

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Table 9.4. Reducing sediment export from urban areas to watercourses Measure

Reference

Sediment traps

Verstraeten et al, (2006)

Swales

Shaw and Kuo (2001), Barrett and Malina (2004), Barrett et al. (2004)

Retention/detention ponds

Hartigan (1988), Schueler et al. (1992)

Street sweeping

CASQA (2003), Schilling (2005)

Curb inlet filters

Smith et al. (2007)

Table 9.5. Estimates of reduction efficiencies of best management practices for urban sediment (Hallock, 2007) Measure

Average efficiency (%)

No. of studies

Catch basin

52

2

Centrifugal separator

12

2

Constructed wetlands

72

13

Dry ponds, vegetated

69

7

Filter strips

64

7

Porous pavement

71

8

Infiltration basins

85

4

Street sweeping

74

2

Wet ponds

56

31

9.2

Monitoring the Effectiveness of Measures

9.3

Muleta (2010) used field measurements, obtained using an extensive range of measures, to calibrate a model, based on the Soil Water Assessment Tool (SWAT), that could be used to predict the effectiveness of such measures over a wider spatial area. Similar modelling was done by Bracmort et al. (2004, 2006). Sediment source tracking can be used to evaluate the effectiveness of measures (Minella et al., 2008). However, the results of specific field investigations can depend on the spatial scale of the study (e.g. Delmas et al., 2012). Barrett and Malina (2004) reviewed the effectiveness of silt fence measures.

The Use of Modelling for the Design and Evaluation of Measures

In the USA, the KINEROS and SWAT models have been used to simulate the effectiveness of soil conservation measures (Arabi et al., 2007; Van Liew et al., 2007). In Europe, the EUROSEM–GRIDSEM modelling system was developed for a similar purpose (Morgan et al., 1994; Botterweg et al., 1998; Morgan et al., 1998; Kinnell, 1999). It was initially evaluated for European catchments (Folly et al., 1999), but has since been more widely used, for example in China (Cai et al., 2005), Kenya (Mati et al., 2006), Costa Rica, Nicaragua and Mexico (Veihe and Al, 2001). It has also been compared with both KINEROS and SWAT (Smith et al., 1995).

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10 Standards and Targets

Table 1.1, at the beginning of this review, lists some of the initial sediment limits, which were based on varying ecological impacts. Collins et al. (2011) listed international sediment targets for river catchment management, while APEM (2007) and Cooper et al. (2008) discussed UK requirements. A special report for the Canadian Council of Ministers gives standards for Canada (CCME, 2001). Foster et al. (2011) used palaeolimnological analyses to estimate sedimentation rates in water bodies prior to agricultural intensification in the mid-20th century to assist with setting sediment standards. Rose et al. (2011) addressed the wider picture for European lakes since AD 1850.

Table 10.1 shows proposed target and critical suspended sediment yields for various catchment types in England and Wales, while Table 10.2 demonstrates the diversity of the approaches taken and the turbidity limits set by a sample of regulations related to different regions. Note, for instance, that Australia and New Zealand differentiate between catchments on the basis of altitude. In addition, New Zealand now has targets for deposited sediment (Clapcott et al., 2011).

Table 10.1. Proposed target and critical suspended sediment yields for various catchment types in England and Wales (Cooper et al., 2008; Collins et al., 2011) Catchment type

Target suspended sediment yield (t/km2 per year) (lower quartile)

Critical suspended sediment yield (t/km2 per year) (upper quartile)

High impermeable peat

50

> 150

Low impermeable peat

No data

No data

Low impermeable (non-peat, non-chalk)

40

> 70

Low impermeable (non-peat, non-chalk)

20

> 50

High impermeable (non-peat, non-chalk)

10

> 20

High permeable (non-peat, non-chalk)

No data

No data

Low chalk

2

> 5

69

70

10–20 NTU 2–15 NTU 1–50 NTU

South western upland and lowland

Tropical upland and lowland

South central upland and lowland

4.1 NTU 5.6 NTU

Upland (> 150- to < 1500-m altitude)

Lowland (< 150-m altitude)

New Zealand:

6–50 NTU

South-eastern lowland (< 150-m altitude)

≤ 10% cumulative increase relative to upstream control

May not exceed 25 NTU above natural

Turbidity

2–25 NTU

Australia:

Australia and New Zealand Guidelines for Freshwater and Marine Water Quality (ANZECC, 2000)

Suspended and settleable solids should not reduce the depth of the compensation point for photosynthetic activity by > 10% from the seasonally established norm for aquatic life

In all other surface waters no sediment loads (suspended or deposited) that can cause adverse effects on aquatic animal or plant life, or their reproduction or habitat may be present

The percentage accumulation of fine sediment between 0.1 and 4.0 mm in the gravel bed of waters used by anadromous or resident fish for spawning may not be increased more than 5% by weight above natural conditions and must not exceed a maximum of 30% by weight (grain size accumulation graph)

< 10% above background when background is > 250 mg/L

< 25 mg/L above background when backgrounds are > 25 but < 250 mg/L

High flow:

< 5 mg/L above background (> 24 hours but < 30 days exposure)

< 25 mg/L above background (< 24 hours exposure)

Low flow:

≤ 25 mg/L (apart from exceptional conditions e.g. floods and droughts)

SSs

South-eastern upland (> 150- to < 1500-m altitude)

Oregon (USA)

Alaska (USA)

Alaska Water Quality Standards (ADEC, 2016)

Oregon Department of Water Quality Standards (Oregon DEQ, 2005)

Canada

Canadian Environmental Quality Guidelines for Protection of Freshwater Aquatic Life (CCME, 2007)

USA

European Union

Freshwater Fish Directive (78/659/EEC and 2004/44/EC)

National Recommended Water Quality Criteria (USEPA, 2007)

Country/state/region

Directive/Regulation

Table 10.2. Examples of standards/regulations for various countries

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Abbreviations

ACP Agricultural Catchments Programme ADCP Acoustic Doppler current profiler As Arsenic AVHRR Advanced very high resolution radiometers AWS Automatic water sampler Combined Automated, Flood, Elevation and Stream Power (system) CAFES Cu Copper Median particle diameter D50 EPA Environmental Protection Agency FTU Formazin turbidity units INCA-Sed Integrated Catchment Model for Sediments IR Infra-red LISST Laser in situ scattering and transmissometry NTU Nephelometric turbidity units OBS Optical backscatter sensor Pb Lead PEEP Photo-electronic erosion pin Q River discharge SDR Sediment delivery ratio SS Suspended solid SSC Suspended sediment concentration SSL Suspended sediment load SSY Suspended sediment yield SWAT Soil Water Assessment Tool TSS Total suspended solid USEPA United States Environmental Protection Agency USGS United States Geological Survey WFD Water Framework Directive WWTP Wastewater treatment plant Zn Zinc

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AN GHNÍOMHAIREACHT UM CHAOMHNÚ COMHSHAOIL Tá an Ghníomhaireacht um Chaomhnú Comhshaoil (GCC) freagrach as an gcomhshaol a chaomhnú agus a fheabhsú mar shócmhainn luachmhar do mhuintir na hÉireann. Táimid tiomanta do dhaoine agus don chomhshaol a chosaint ó éifeachtaí díobhálacha na radaíochta agus an truaillithe.

Monatóireacht, Anailís agus Tuairisciú ar an gComhshaol • Monatóireacht a dhéanamh ar cháilíocht an aeir agus Treoir an AE maidir le hAer Glan don Eoraip (CAFÉ) a chur chun feidhme.

• Tuairisciú neamhspleách le cabhrú le cinnteoireacht an rialtais náisiúnta agus na n-údarás áitiúil (m.sh. tuairisciú tréimhsiúil ar staid Chomhshaol na hÉireann agus Tuarascálacha ar Tháscairí).

Is féidir obair na Gníomhaireachta a roinnt ina trí phríomhréimse: Rialú: Déanaimid córais éifeachtacha rialaithe agus comhlíonta comhshaoil a chur i bhfeidhm chun torthaí maithe comhshaoil a sholáthar agus chun díriú orthu siúd nach gcloíonn leis na córais sin.

Rialú Astaíochtaí na nGás Ceaptha Teasa in Éirinn • Fardail agus réamh-mheastacháin na hÉireann maidir le gáis cheaptha teasa a ullmhú.

• An Treoir maidir le Trádáil Astaíochtaí a chur chun feidhme i gcomhair

Eolas: Soláthraímid sonraí, faisnéis agus measúnú comhshaoil atá ar ardchaighdeán, spriocdhírithe agus tráthúil chun bonn eolais a chur faoin gcinnteoireacht ar gach leibhéal.

breis agus 100 de na táirgeoirí dé-ocsaíde carbóin is mó in Éirinn.

Taighde agus Forbairt Comhshaoil • Taighde comhshaoil a chistiú chun brúnna a shainaithint, bonn eolais a chur faoi bheartais, agus réitigh a sholáthar i réimsí na haeráide, an uisce agus na hinbhuanaitheachta.

Tacaíocht: Bímid ag saothrú i gcomhar le grúpaí eile chun tacú le comhshaol atá glan, táirgiúil agus cosanta go maith, agus le hiompar a chuirfidh le comhshaol inbhuanaithe.

Measúnacht Straitéiseach Timpeallachta • Measúnacht a dhéanamh ar thionchar pleananna agus clár beartaithe ar an

Ár bhFreagrachtaí

gcomhshaol in Éirinn (m.sh. mórphleananna forbartha).

Ceadúnú

Cosaint Raideolaíoch

Déanaimid na gníomhaíochtaí seo a leanas a rialú ionas nach ndéanann siad dochar do shláinte an phobail ná don chomhshaol:

• Monatóireacht a dhéanamh ar leibhéil radaíochta, measúnacht a

• saoráidí dramhaíola (m.sh. láithreáin líonta talún, loisceoirí, stáisiúin

• Cabhrú le pleananna náisiúnta a fhorbairt le haghaidh éigeandálaí ag

dhéanamh ar nochtadh mhuintir na hÉireann don radaíocht ianúcháin.

aistrithe dramhaíola);

• gníomhaíochtaí tionsclaíocha ar scála mór (m.sh. déantúsaíocht

eascairt as taismí núicléacha.

• Monatóireacht a dhéanamh ar fhorbairtí thar lear a bhaineann le saoráidí

cógaisíochta, déantúsaíocht stroighne, stáisiúin chumhachta);

• an diantalmhaíocht (m.sh. muca, éanlaith); • úsáid shrianta agus scaoileadh rialaithe Orgánach Géinmhodhnaithe

núicléacha agus leis an tsábháilteacht raideolaíochta.

• Sainseirbhísí cosanta ar an radaíocht a sholáthar, nó maoirsiú a dhéanamh ar sholáthar na seirbhísí sin.

(OGM);

• foinsí radaíochta ianúcháin (m.sh. trealamh x-gha agus radaiteiripe, foinsí tionsclaíocha);

• áiseanna móra stórála peitril; • scardadh dramhuisce; • gníomhaíochtaí dumpála ar farraige.

Forfheidhmiú Náisiúnta i leith Cúrsaí Comhshaoil • Clár náisiúnta iniúchtaí agus cigireachtaí a dhéanamh gach bliain ar shaoráidí a bhfuil ceadúnas ón nGníomhaireacht acu.

• Maoirseacht a dhéanamh ar fhreagrachtaí cosanta

Treoir, Faisnéis Inrochtana agus Oideachas • Comhairle agus treoir a chur ar fáil d’earnáil na tionsclaíochta agus don phobal maidir le hábhair a bhaineann le caomhnú an chomhshaoil agus leis an gcosaint raideolaíoch.

• Faisnéis thráthúil ar an gcomhshaol ar a bhfuil fáil éasca a chur ar fáil chun rannpháirtíocht an phobail a spreagadh sa chinnteoireacht i ndáil leis an gcomhshaol (m.sh. Timpeall an Tí, léarscáileanna radóin).

• Comhairle a chur ar fáil don Rialtas maidir le hábhair a bhaineann leis an tsábháilteacht raideolaíoch agus le cúrsaí práinnfhreagartha.

• Plean Náisiúnta Bainistíochta Dramhaíola Guaisí a fhorbairt chun dramhaíl ghuaiseach a chosc agus a bhainistiú.

comhshaoil na n-údarás áitiúil.

• Caighdeán an uisce óil, arna sholáthar ag soláthraithe uisce phoiblí, a mhaoirsiú. •Obair le húdaráis áitiúla agus le gníomhaireachtaí eile chun dul i ngleic le coireanna comhshaoil trí chomhordú a dhéanamh ar líonra forfheidhmiúcháin náisiúnta, trí dhíriú ar chiontóirí, agus trí mhaoirsiú a dhéanamh ar leasúchán.

Múscailt Feasachta agus Athrú Iompraíochta • Feasacht chomhshaoil níos fearr a ghiniúint agus dul i bhfeidhm ar athrú iompraíochta dearfach trí thacú le gnóthais, le pobail agus le teaghlaigh a bheith níos éifeachtúla ar acmhainní.

• Tástáil le haghaidh radóin a chur chun cinn i dtithe agus in ionaid oibre, agus gníomhartha leasúcháin a spreagadh nuair is gá.

• Cur i bhfeidhm rialachán ar nós na Rialachán um Dhramhthrealamh Leictreach agus Leictreonach (DTLL), um Shrian ar Shubstaintí Guaiseacha agus na Rialachán um rialú ar shubstaintí a ídíonn an ciseal ózóin.

• An dlí a chur orthu siúd a bhriseann dlí an chomhshaoil agus a dhéanann dochar don chomhshaol.

Bainistíocht Uisce • Monatóireacht agus tuairisciú a dhéanamh ar cháilíocht aibhneacha, lochanna, uiscí idirchriosacha agus cósta na hÉireann, agus screamhuiscí; leibhéil uisce agus sruthanna aibhneacha a thomhas.

• Comhordú náisiúnta agus maoirsiú a dhéanamh ar an gCreat-Treoir Uisce.

• Monatóireacht agus tuairisciú a dhéanamh ar Cháilíocht an Uisce Snámha.

Bainistíocht agus struchtúr na Gníomhaireachta um Chaomhnú Comhshaoil Tá an ghníomhaíocht á bainistiú ag Bord lánaimseartha, ar a bhfuil Ard-Stiúrthóir agus cúigear Stiúrthóirí. Déantar an obair ar fud cúig cinn d’Oifigí:

• • • • •

An Oifig um Inmharthanacht Comhshaoil An Oifig Forfheidhmithe i leith cúrsaí Comhshaoil An Oifig um Fianaise is Measúnú An Oifig um Cosaint Raideolaíoch An Oifig Cumarsáide agus Seirbhísí Corparáideacha

Tá Coiste Comhairleach ag an nGníomhaireacht le cabhrú léi. Tá dáréag comhaltaí air agus tagann siad le chéile go rialta le plé a dhéanamh ar ábhair imní agus le comhairle a chur ar an mBord.

EPA Research Report 176

SILTFLUX Literature Review Authors: D. Lawler, A. Rymszewicz, L. Conroy, LT. O’Sullivan, M. Bruen, J. Turner, M. Kelly-Quinn

Identifying Pressures Fine sediment delivery to rivers is increasingly recognised internationally as a substantial water quality and hydro-ecological problem. The SILTFLUX project aimed to improve knowledge of fine sediment delivery as a pollution pressure in Irish rivers. The project has studied sediment flux dynamics with respect to key flow events and their actual and potential ecological impacts in different Irish river systems that are subject to variable land-use pressures. The SILTFLUX Literature Review distils the current knowledge of these effects and of methodologies for reducing their impacts in conditions typical of Ireland.

Informing Policy The SILTFLUX project will help inform environmental management and policy in the setting of standards for suspended sediment fluxes and concentrations appropriate for the protection of sensitive catchments in Ireland. The SILTFLUX Literature Review synthesises the considerable international debate that has surrounded the basis for establishing such standards, particularly on the issues of (i) whether they should be based on sediment loads, suspended sediment concentrations, deposited sediment or all three, and (ii) how such standards can account for the biological impacts of both transported and deposited sediment. Addressing these issues in a framework that is easily and reliably measurable and which lends itself to monitoring on a national scale remains a challenge.

Developing Solutions The SILTFLUX project has also identified the benefits of suspended sediment reduction possible from a broad range of measures and land management practices, to support the development of mitigation policies. A review of the published literature and existing measured data was undertaken to establish an initial “state of the art” position. The SILTFLUX Literature Review was international in scope but included key foci on information that was relevant to the conditions and pressures experienced in Irish catchments. Existing sediment datasets from the literature, as well as the projects own data, were compiled in a database that identifies the range of fluxes to be expected in Irish river systems.

EPA Research: McCumiskey House, Richiew, Clonskeagh, Dublin 14. Phone: 01 268 0100 Twitter: @EPAResearchNews Email: [email protected]

www.epa.ie EPA Research Webpages www.epa.ie/researchandeducation/research/

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SILTFLUX Literature Review - EPA

Report No. 176 SILTFLUX Literature Review Authors: D. Lawler, A. Rymszewicz, L. Conroy, J. O’Sullivan, M. Bruen, J. Turner, M. Kelly-Quinn www.epa.i...

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