1 ABSTRACT

Coastal erosion is to some extent affecting all European coasts with current

estimates indicating that 15,100 km of the coastline (20%) is actively retreating and

about 15 km2 of land is being lost or seriously impacted by erosion processes on an

annual basis. This widespread erosion is due to a combination of both natural

processes and anthropogenic factors. The cost of mitigation action towards

protecting the coastline and its assets against the risk of erosion and flooding is

predicted to rise to €5,400 million per year between 1990 and 2020.

In recent years, the value of intertidal habitats in coastal defence has been

recognised and the focus in coastal management has undergone a paradigm shift

from hard engineered defences to soft engineered defences through recreating or

restoring such habitats. Managed realignment techniques have been successful but it

became apparent that surface elevation played a critical role in the establishment of

saltmarsh vegetation which led to the question of whether it would be possible to

facilitate plant colonization of a site by raising the level through foreshore recharge

using dredged material.

This study evaluates the potential of using muddy dredged material to recharge

intertidal areas in the Orwell estuary measurable in terms of ecological

enhancement and flood defence value. The trial site was created at Shotley Point in

1997 and the most recent topographic and vegetation survey was conducted in 2006.

Results suggest that the trial has been successful in increasing the elevation of the

site to a level that will allow plant colonization and the site is now providing 32 m

width of saltmarsh protection to the eroding sea wall, which in terms of coastal

defence can be seen as a success. However, the conservation value of the site is

questionable and further monitoring is required before it can be said with certainty

that this method of restoring habitats can be used as a means of recreating habitats

of larger areas.

1

2 INTRODUCTION

2.1 Coastal erosion and habitat loss

Coastal erosion is to some extent affecting all European coasts with current

estimates indicating that 15,100 km of the coastline (20%) is actively retreating and

about 15 km2 of land is being lost or seriously impacted by erosion processes on an

annual basis (Doody et al. 2004). This widespread erosion is due to a combination

of both natural processes e.g. sea level rise, winds and storms; and anthropogenic

factors e.g. coastal engineering and land reclamation (Doody et al. 2004).

The cost of mitigation action towards protecting the coastline and its assets against

the risk of erosion and flooding are ever increasing. In 2001, public expenditure

dedicated to protecting European coasts was estimated to be in the order of €3,200

million compared to €2,500 million in 1986. Previous studies for the UN-IPCC

predicted the average cost of coastal erosion would rise to €5,400 million per year

between 1990 and 2020 (Doody et al. 2004).

A survey monitored by the Joint Nature Conservation Committee (JNCC) estimated

that Great Britain has approximately 10,000 km of coastline (French 1997, 68),

much of which has been heavily modified through reclamation and construction of

hard engineered structures i.e. sea walls, to protect the generally low lying

reclaimed land. Evidence for land claim in the UK dates back to Roman Times in

areas such as the Severn Estuary with extensive reclamation around the Wash from

the 17th Century onwards (Healy & Hickey 2002). The onset of the Industrial

Revolution in the 19th Century led to further land claim in estuaries for port

development (French 1997, 58). At present there are 860 km of soft cliffs (23% of

the coastline) protected by erosion defence structures and in excess of 1,259 km of

flood defences protecting 2,347 km2 of former flood plains (Crooks et al. 2004). It

would appear that at the time of construction little thought was given to the long

2

term consequences of permanently fixing in place what is essentially a dynamic

structure which is constantly evolving in response to natural forcing factors.

Most sea walls are fronted by intertidal habitats which act as natural buffers

between marine and terrestrial environments. Under natural conditions, a rise in sea

level would result in these habitats trying to regain their dynamic equilibrium by

relocating inland through lateral and vertical translation (French 2001, 41). This

movement of a coastal form is achieved by initial erosion of the habitat profile

which will be reformed higher up in the tidal frame, commonly referred to as the

Bruun rule (French 2001, 41). However, the presence of a sea wall prevents this

natural transgression resulting in the habitats being squeezed between the fixed sea

defence and elevated sea level – a process known as coastal squeeze (Pethick 2001)

- the consequences of which are a net decline in intertidal habitats. In the face of

climate change and projected sea level rises of 0.05 to 0.32 m between 1990 and

2050 (IPCC 2001), more and more of this habitat will be lost unless action is taken

to alleviate the problem, halt erosion and reverse the trend.

In the UK, the area most at risk from rising sea levels is the south-east of England

where sea level is predicted to rise by 19 to 79cm by 2080 (UKCIP 2002). In

addition to eustatic sea level rise, southern Britain is experiencing land level

adjustments associated with isostatic rebound following the last ice age which

contributes further to the rise in relative sea level (Boorman 2003). For this reason,

extensive research has been focused on the coastline of Essex in an attempt to

protect the land through the restoration of intertidal habitats (Dixon et al. 1998 &

Cooper et al. 2001). Over 70% of sea walls in Essex are fringed by saltmarshes

which offer protection by reducing wave energy and thereby reducing the impact of

wave action on sea walls (Tabor 2003). Coastal saltmarshes can be defined as

environments found high in the intertidal zone composed of fine sediment which

support halophytic plants and are periodically flooded as a result of fluctuations

(tidal or non-tidal) in the level of the adjacent water body (Adam 1990, 1). They are

normally fronted by mudflats or sandflats and increase in elevation by moving

3

upwards and landwards into freshwater environments (Allen & Pye 1992), forming

distinct zones as elevation increases: pioneer; low marsh; middle marsh; and high

marsh; with each zone experiencing a different tidal regime depending on its

elevation. These zones are characterized by the plant species that dominate the area

which in turn will be present or absent in response to the frequency of tidal

inundation (Boorman 2003).

Over the last 400 years, large areas of saltmarsh have been reclaimed and converted

to farmland (Pye 1992). Saltmarshes make up approximately 20% of the coastline

of England and Wales and being tide dominated environments, are mainly found in

or close to estuaries (Brampton 1992). The total area covered by saltmarsh in Great

Britain is approximately 44,730 ha (Allen & Pye 1992) most of which occurs in the

east and south-east of England. The Thames estuary, the Essex coast, Liverpool and

Morecambe Bay and the Wash account for 60% of the UK total (Tabor 2003). The

last century has seen large areas of saltmarsh being lost through coastal squeeze and

67% of the eastern coastline has shown landward retreat (Taylor et al. 2004).

Saltmarshes in the south-east of England have been eroding rapidly since 1960 and

has become a major concern for conservation and coastal managers (Wolters et al.

2005). A study by Cooper et al (2001) on saltmarsh erosion in Essex estimated that

over a 25 year period between 1973 and 1998 1,000 ha of saltmarsh were lost

through erosion equating to an average rate of loss of 40 ha year-1. Due to the

transient nature of saltmarshes, it is almost impossible to say how much of this

valuable habitat has been lost in total as, even in the absence of anthropogenic

pressure, they would accrete and erode in response to natural processes such as

changing sea levels and sediment availability. However, Davidson et al (1991)

proposed that the total loss of saltmarsh in the UK from all causes stands at about

91,250 ha since Roman times.

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2.2 Value of saltmarsh in flood defence

Saltmarshes are valuable in terms of flood defence as they are natural dissipaters of

tidal currents and waves (Moller & Spencer 2002). They act by reducing wave

heights and energy reaching any existing hard engineered flood defence structures

thereby reducing erosion and maintenance costs. Despite the recognition of the

effectiveness of saltmarsh vegetation in dissipating wave energy and their value in

coastal defence, there have been very few quantitative field studies focussed on the

subject, although laboratory simulations have been verified by field studies in the

US namely by Wayne (1976) and Knutson et al (1982). The findings from both

these studies were fairly similar and reported reductions in wave energy of up to

92% and in some cases almost 100% along 20-30 m long transects of marsh habitat

comprised of tall dense Spartina alterniflora. However, these results cannot be

applied directly to the behaviour and value of saltmarshes in the UK as there are

fundamental physical differences between the two habitats. The vegetation cover in

the two studies mentioned above are markedly different to the mixed marsh

communities found in north-west Europe (Moller et al. 1999). In addition to this,

tidal ranges and sediment characteristics between the two regions are also very

different. Sediments that form UK marshes are mainly clastic and the average tidal

range is 3-12 m, whereas marsh sediments in the US consist largely of peat and the

habitats exist in much lower tidal ranges, often less than 2 m (Atkinson et al. 2001).

Studies on this topic in the UK has mainly been restricted to physical scale model

experiments (Hydraulics Research 1980) which reported wave height reductions of

49% across an 80 m width of saltmarsh. These findings were further supported by

Moller et al (1999) who compared field measurements of wave energy

transformations at a site at Stiffkey in north Norfolk with one dimensional model

calculations based on theoretical knowledge of wave dissipation and reported an

average reduction in wave height of 63% over 200 m saltmarsh.

The ability of saltmarshes to reduce wave height can be translated into monetary

terms through the lower maintenance costs of existing sea walls. A report by the

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National Rivers Authority (NRA, now the Environment Agency) in 1992 stated that

an 80 m strip of saltmarsh in front of a sea wall would reduce the maintenance cost

to £1 m-1 as opposed to £50 m-1 in the absence of any saltmarsh (Table 1).

Width of saltmarsh/m Wall height/m

Cost of new wall/£m-1

Maintenance cost/£m-1

80 3 400 1 60 4 500 5 30 5 800 15 6 6 1500 25-30 0 12 3000-5000 50

Table 1: Comparison of capitol and maintenance costs of sea walls with varying amounts of

saltmarsh protection (NRA 1992).

Capitol costs of building new sea walls are also reduced as the height of the sea wall

required decreases with increasing saltmarsh width (Table 1). King & Lester (1995)

calculated that a complete loss of saltmarsh in Essex would incur minimum costs of

over ₤600 million for rebuilding of sea walls highlighting the fact that saltmarshes

are a valuable economic asset.

As our understanding of coastal processes has improved, so has our approach to

coastal management. It is now recognized that the coastline is a dynamic

environment that has to have the flexibility to respond to perturbations via sediment

movements and natural evolution. Traditional ‘hard’ engineering techniques which

saw the use of concrete and boulders to permanently ‘fix’ the coastline in place may

have offered protection in the short term but in the long term has proven to be

detrimental to the longevity and sustainability of such techniques.

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2.3 Conservation value of saltmarsh

Aside from their importance and economic value in coastal defence, saltmarshes

have also been recognized as areas of high biodiversity for many years and are

considered to be habitats of international importance for conservation.

Approximately 80% of the area of saltmarshes in the UK have been identified as

Sites of Scientific Interest (SSSI) and 10% have been designated as Special

Protection Areas (SPA) under the EC Birds Directive 70/409/EEC, or Special Areas

of Conservation (SAC) under the EC Habitats Directive 92/43/EEC (Cooper et al.

2001). They provide feeding, breeding and roosting areas for a plethora of bird

species especially where unimproved meadowland has been replaced with

agricultural land. Additionally, they are areas of high productivity and provide a

source of organic matter and nutrients for fish and invertebrates (Boorman 2003).

The EC Habitats Directive promotes a ‘no-net-loss’ policy on all Natura 2000 sites

(collective name for SPAs and SACs across Europe) and was transposed into British

law by the Conservation Regulations in 1994 (Ledoux et al. 2000). In terms of

protection of saltmarshes the Directive proposed four main principles to implement

this policy, namely through: no further loss of wetlands, no further degradation of

wetlands, wise use of wetlands, and wetland improvement and restoration (ABP

1998). However, it should be noted that this is a blanket policy which applies to

sites across Europe and detailed implementation is passed down to a national and

local level. Nationally, the nature conservation authorities in the UK (Natural

England, Scottish Natural Heritage and Countryside Council for Wales) have all

adopted the no-net-loss policy and have set out requirements for developers to

recreate and restore wetlands to compensate for habitat loss and such mitigation

schemes are becoming increasingly common. On a local level, the regulations are

implemented through the development of Coastal Habitat Management Plans

(ChaMPs) and Shoreline Management Plans (SMPs). The UK Biodiversity Action

Plan also lays down objectives and targets for saltmarsh habitats and aims to

recreate 100 ha year-1 to ensure no further loss is experienced (UKBAP 1999).

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2.4 Habitat recreation schemes

With the recognition of the economic and ecological value of saltmarsh, the focus in

coastal management has shifted from ‘hard’ engineered structures to a ‘soft’

engineered approach of recreating or restoring intertidal habitats on both the

landward and seaward faces of fixed defence lines. In recent years, managed

realignment of coastal defences became a popular technique but success from trial

schemes have been varied. The first small scale trial (0.8 ha) was conducted in 1991

at Northey Island in the Blackwater Estuary and was successful in establishing new

saltmarsh habitat behind the breached sea wall (Dagley 1995). A larger scale trial

(21 ha) was then conducted at Tollesbury in Essex which was also successful in

creating new saltmarsh (Tabor 2003). Subsequent schemes following on from these

trials were not quite as successful in achieving vegetation cover and it became clear

that the land behind the breached sea wall had to be of a sufficient elevation for

successful vegetation colonization (Boorman 2003). Although other factors play a

role in the success of such schemes such as: tidal prism; estuarine morphology; tidal

hydraulics; site history; surface gradient; sediment characteristics; accretion

processes; and wave climate (Burd 1995), numerous studies have shown that

surface elevation relative to the tidal frame is the key factor influencing vegetation

establishment on saltmarshes (Davy 2000). This development in the understanding

of saltmarsh recreation led to the question of whether it would be possible to

facilitate plant colonization of a site by raising the level through foreshore recharge

using dredged material.

2.5 Foreshore recharge and beneficial use of dredged material

On an annual basis, up to 50x106 tonnes of material is dredged from coastal areas in

the UK during routine maintenance and capitol dredging (Fletcher et al. 2001).

Maintenance dredging is an essential activity that is required to keep navigation

channels clear and safe for use. Since the early 1990’s there has been a move

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towards considering coastal sediments and dredged material from ports and

harbours as a potential resource before classifying it as marine waste which requires

offshore disposal (Fletcher et al. 2001). Under the Food and Environment

Protection Act 1985 (FEPA), the Department for Environment, Food and Rural

Affairs (DEFRA, formerly MAFF) is responsible for granting licenses for disposal

of such material and have developed a policy whereby applicants are required to

show they have given due consideration to the alternative uses of the material

before applying for such a license (Colenutt 1999). In addition to this policy, the

UK is a signatory to a number of international conventions which regulate the

dumping of noxious substances at sea i.e. the London Convention 1972 and the

OSPAR Convention 1992, and therefore has a duty to recycle dredged material

(MAFF 1999). Most of the dredged material is of an unpolluted muddy nature

which is not suitable for use in engineering and construction projects but can be

used in habitat recreation schemes. Utilising the material in this way meets three

main aims of the above Conventions and the aforementioned environmental

legislations: beneficial use; maintain or increase biodiversity; and support coastal

defences in a more integrated and sustainable manner (Fletcher et al. 2001).

Foreshore recharge and beneficial use of dredged material is not a new concept and

has been used in the United States for the past 30 years to recreate marshes with

reports of over 16,000 ha of riparian and coastal wetlands being restored in this

manner (PIANC 1992). However, the large number of restoration schemes

undertaken have not been without criticism, mainly due to the low success or failure

of some schemes implemented in the 1980s as a result of poor planning and unclear

objectives (ABP Research 1998). Nevertheless, trend data for most of the work

suggests that well planned and executed projects can be highly successful in

restoring habitats which are favourably comparable with natural sites (Landlin et al.

1995). A primary objective of these restoration schemes was to create self

sustaining sites which display continued natural ecological succession, a

development which has been demonstrated by all the man made sites that have been

considered to be successful (Landlin 1995). In some instances, the natural evolution

9

of the site has resulted in the evolution of habitats different to those originally

proposed, although this does not necessarily indicate the failure of the scheme

(Landlin 1995). The US Army Corps of Engineers (USACE) who have been using

dredged material to create Spartina alterniflora marshes since 1969, have developed

a strategic framework for the selection of suitable habitat development schemes in

North America (Streever 2000). The lessons learnt from the thousands of restoration

schemes undertaken over the years by the United States, both successful and

unsuccessful, have served to identify elements which are important to the success of

a restoration project such as careful planning, long-term monitoring and good

understanding of the physical and ecological function of the habitats to be restored

(ABP Research 1998).

In comparison, there are much fewer studies of this nature occurring the UK

although there are small scale projects (up to 20x103 m3 per scheme) trialing this

technique at sites along the Suffolk and Essex coast. These sites created in 1998

include: north Shotley Marsh, Trimley Marsh, Horsey Island, Old Hall Point,

Tollesbury Wick and Wallasea Ness. The sediments required to recharge the above

sites are derived from channel excavations through Harwich Haven to the port of

Felixstowe (Atkinson et al.2001). The availability of the necessary large volumes of

appropriate sediment that is required to successfully recharge a site could limit the

effectiveness of this technique. In the south-east of England, approximately 3x106

m3 of relatively pollutant free fine sediment are raised through annual maintenance

dredging by Harwich Haven Authority (French & Watson 1999). If the trial

schemes prove to be successful, this beneficial use of dredged material could be

used to support managed realignment in order to achieve the recreation targets set

by the Habitats Directive (1994) to replace past and predicted habitat loss.

As already mentioned above, many estuaries where most dredging activity takes

place are designated Natura 2000 sites and environmental legislation dictates which

activities are acceptable within such areas. It therefore stands to reason that if the

beneficial uses of dredged material are to be promoted, it will require assistance

10

from the environmental bodies and local authorities that are responsible for

implementing the Habitats Regulations. The marrying of all the aforementioned

policies will hopefully move towards an ever more holistic and sustainable view to

managing the British coastline.

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3 STUDY SETTING

3.1 Aim and objectives

The aim of this study is to assess the foreshore recharge trial at Shotley Point in the

Orwell estuary and investigate the geomorphological development of the site

compared with previous surveys by French & Watson (1999 and 2000) and more

recently Rossi (2002).

Specific objectives of this study are:

• To conduct a literature review on historical saltmarsh loss and determine the

possible causes of erosion;

• To determine changes in elevation levels of the recharge site and the crest level

and width of the gravel bund;

• To assess profile evolution of the recharge site;

• To assess the development of saltmarsh vegetation;

• To assess the suitability of using muddy dredged material for the development

of saltmarsh habitat both within the Orwell estuary and at other sites;

3.2 Study site

Shotley Point is located at the mouth of the Orwell Estuary in the county of

Suffolk. Of the 44,730 ha of saltmarsh found in the UK (Allen & Pye 1992), 2%

of this is found within the Suffolk estuaries (Beardall et al. 1988). Large areas of

high marsh in this area were reclaimed by direct enclosure behind sea walls

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between the 11th-13th and 16th-17th century although the earliest embankments

can be dated back to Roman times. The total area of reclaimed marsh and

mudflat in the Suffolk estuaries is estimated to be 10,817 ha or 2.5 times the

area of the intertidal zone that remains today (Beardall et al. 1988).

The Orwell Estuary has a tidal length of 18 km and is mesotidal with spring

tides ranging from 3.6 m at Harwich to 3.9 m at Ipswich (Guthrie & Cottle

2002). The tidal range at Shotley Point would be similar to that of Felixstowe

which is 3.4 m although there may be some minor differences as Felixstowe is

an open coast gauge and not an inner estuary. Mean tide levels for Felixstowe in

2006 are shown in Table 2 below:

Mark

Chart Datum/m

(CD)

Ordnance Datum/m

(OD)

Mean High Water Spring

(MHWS) 3.80 1.85

Mean High Water Neap

(MHWN) 3.10 1.15

Mean Low Water Spring

(MLWS) 0.40 -1.55

Mean Low Water Neap

(MLWN) 1.00 -0.95

Highest Astronomical

Tide (HAT) 4.20 2.25

Lowest Astronomical Tide

(LAT) -0.10 -2.05

Mean Sea Level

(MSL) 2.10 0.15

Table 2: Felixstowe tide data (Admiralty Tide Tables 2006).

13

14

Three major dredging schemes since 1947 have removed 4x106 m3 of silt and

routine maintenance dredging removes a further 3.8x104 m3 year-1. The channel

is now dredged to a depth of 5.8 m which has increased the volume of the

estuary by 56%.

Cooper et al. (2001) indicated that the River Orwell has suffered a net loss of

23% of its original area of saltmarsh between 1988 and 1997. Consequently, the

sea walls have suffered serious toe erosion and are in a poor state of repair. A

sediment monitoring study in the Orwell estuary undertaken by the Suffolk

Wildlife Trust (1991) revealed a drop in surface elevation of the upper mudflats

at Shotley Point of up to 0.93 m although the surfaces of Shotley Marshes in the

same survey appeared to have accreted by 0.24 to 0.40 m.

The trial placement north of Shotley Marina (Figure 1) was initiated in

December 1997 and was originally instigated as a joint venture between

Harwich Haven Authority (HHA) and the Environment Agency (EA).

Approximately 13x103 m3 of gravel, dredged from the Harwich approach

channel, was used to construct a protective bund to enclose an intertidal area of

430x700 m in front of the degrading sea wall. This enclosure was then filled

with 22x103 m3 of maintenance dredged silt from the Trinity berths at

Felixstowe which is less than half a mile from Shotley (Fletcher et al. 2001).

Monitoring work was conducted over three years by UCL and the results

indicated the positive development of the recharge into mudflat habitat. The site

was topped up with a further 10.7x103 m3 of silt in December 2000 to increase

the elevation again as the initial increase had fallen due to the natural

dewatering process of the sediment. Monitoring for the HHA officially ceased in

2000 but the site was re-evaluated by Rossi in August 2001. Between the

surveys in 2001 and 2006, HHA undertook a further top-up recharge, although

there are no documented reports of how much dredged material was deposited

on the site or when this occurred (French J., personal communication 2006).

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Figure 1: Location of Shotley Recharge Site (French & Watson 1999).

4 METHODOLOGY

4.1 Topographic survey

Topographic surveys of the recharged site were taken on the 12th June 2006 using an

electronic ‘total station’ (Topcon). Critical height control and reduction to Ordnance

Datum (OD) were carried out by levelling. All surface elevation data in this survey are

referenced to OD using EA Survey mark OR6S which is located adjacent to the site and

are given relative to OD which is approximately 1.95 m above local Chart Datum (CD)

at Felixstowe (Admiralty Tides Tables 2006).

Eight of the original nine transects (T1-T8) were re-surveyed at approximate locations

(Figure 11). The first transect from all previous surveys was not re-surveyed as doing so

would have disturbed a pair of Oyster Catchers nesting on the line. No data was

collected further north of the site beyond T8 (past grid reference 234860) as the mud

surface had not compacted enough to allow safe access across the width of the site.

Surface elevation data were recorded along the transects which crossed the width of the

site and extended from the base of the sea wall across the gravel bund and up to the

mean low water mark. Care was taken to ensure data was collected at points which

would record the topography of important features e.g. inner edge, crest, and outer edge

of gravel bund.

Extra data were also collected to record the horizontal position of the gravel bund and a

small dataset recording the presence or absence of saltmarsh vegetation was collected to

provide a general idea of halophytic colonization and their relative elevation levels.

4.2 Analysis of the sample data

The survey data were corrected for rotational error and ordered by Dr. Helene

Burningham which were exported into ArcGIS 9.2.

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Data from previous studies by French & Watson (January 1998, January 1999, January

2000) and Rossi (August 2001) were provided (hereinafter referred to by year only) by

Dr. Helene Burningham and Dr. Jon French, which were also exported into ArcGIS 9.2

to enable comparison of changes in surface elevation within the recharge site and

migration of the gravel bund between 1998 and 2006.

In order to compare the elevation of the recharge site at each sample year the datasets

were interpolated to create a continuous surface using the measurements made at point

locations across the site to predict the elevation at points which lacked sample data. In

this way, elevation data was extracted for each survey year across T1-T8 at 2 m

intervals.

The landward migration of the gravel bund was mapped along the inner edge using a

combination of the sample point data, elevation maps, and contour maps (Figure 7), to

calculate the horizontal landward migration of the bund since the beginning of the trial.

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5 RESULTS

5.1 Topography

Due to the onshore migration of the gravel bund (this will be discussed in further

detail in section 5.3) the survey site can now be divided into two sections and will

be described from here on as: the inner shore (landward of the bund) and the outer

shore (seaward of the bund).

A simple comparison of the net gain and loss in elevation across the site between

the survey years reveals most of the site experienced a loss in elevation between

1998 and 1999 after the initial recharge due to de-watering processes and auto-

compaction (Figure 2). French & Watson (1999) reported that most of this loss in

elevation was experience between January and May 1998. After this predicted loss

in elevation, the general trend has been one of net gain in elevation in the inner

shore but a net loss in elevation in the outer shore between 1999 and 2006. This is

illustrated by Figures 3-6 which shows the inner edge of the gravel bund for each

survey year roughly dividing the site into the inner and outer shore and the net gains

and losses in elevation on either side.

Detailed comparisons of the changes in elevation are illustrated by Figures 7-11.

The elevation ranges for each year are summarized in Table 3. The elevation ranges

recorded in 1998 are not comparable with other years as there were errors in the

dataset provided. Comparison of the elevation ranges between 2001 and 2006 across

the whole recharge site show that the elevation of the site has increased by 0.02 m at

its highest point in the inner shore but decreased by 2.47 m at its lowest point in the

outer shore.

There was an average gain in elevation of the inner shore of 0.05 m between 2001

and 2006, although this gain was not uniform across the site as is evident by the

increase in elevation range for this period by 0.06-0.43 m (Table 4).

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Elevation Range/m OD Year Lowest Highest

Jan 1998* -1.00 5.74Jan 1999 -1.62 2.91Jan 2000 -1.01 2.44

Aug 2001 -1.03 2.55 Jun 2006 -1.44 2.57

Table 3: Summary of elevation ranges 1998-2006.

* The January 1998 dataset contained several anomalies which have distorted the elevation range.

French & Watson (1999) reported initial recharged mud elevations of 0.9 - 1.3 m OD in the first

survey in January 1998.

The average gain in elevation of the inner shore between 1998 and 2006 was

calculated to be 0.62 m while the average elevation range has increased by 0.38-

0.89 m (Table 4) (based on reported ranges in French & Watson 1999).

Elevation Range/m OD Year Inner Shore Outer Shore

1998 0.90 – 1.30 (-1.01) – 0.25 1999 0.40 – 1.41 (-1.62) – 0.90 2000 0.53 – 1.30 (-1.02) – 0.91 2001 0.85 – 2.13 (-1.07) – 1.49 2006 1.28 – 2.19 (-1.44) – 1.28

Table 4: Elevation ranges for recharge site between 1998 and 2006.

In contrast, comparison of the average elevation range of the outer shore between

2001 and 2006 shows the elevation of this area has decreased by 0.21 m at the inner

edge (closest to the outer edge of the gravel bund in 2006) and 0.37 m at the outer

edge (near the mean low water mark).

However, comparison of the average elevation range of the outer shore between

1998 and 2006 shows an increase at the inner edge of 1.03 m although the average

19

elevation of the outer shore has decreased by 0.43 m at it’s lowest point (near the

mean low water mark). This increase in elevation at the inner edge is due to

increasing areas of the elevated inner shore being exposed to the seaward edge as

the bund moves progressively further inland (Figure 6).

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Figure 2: Net gain/loss in elevation of recharge site between 1998 and 1999.

21

Figure 3: Net gain/loss in elevation of recharge site between 1999 and 2000.

22

Figure 4: Net gain/loss in elevation of recharge site between 2000 and 2001.

23

Figure 5: Net gain/loss in elevation of recharge site between 2001 and 2006.

24

Figure 6: Net gain/loss in elevation of recharge site between 1998 and 2006.

25

Figure 7: Elevation (m OD) of recharge site January 1998.

26

Figure 8: Elevation (m OD) of recharge site January 1999.

27

Figure 9: Elevation (m OD) of recharge site January 2000.

28

Figure 10: Elevation (m OD) of recharge site August 2001.

29

Figure 11: Elevation (m OD) of recharge site June 2006 showing transects T1-T8.

30

5.2 Transect profiles

Figures 12-19 show elevation profiles for each survey year plotted against MHWN

and MHWS tides.

With the exception of T1, trends in the surface elevation in 2006 show the form of

the foreshore is characterized by a flat or seaward slope of the inner shore towards

the gravel bund which peaks sharply followed by a steep seaward grading outer

shore. The elevation profiles of the site for 2006 are very similar to that of previous

surveys and the site can clearly be seen to be progressively shifting landwards as the

gravel bund migrates.

The average elevation of the inner shore is now 0.67 m above the MHWN tide level

at 1.82 m OD ± 0.19 SD compared with 1.78 m OD ± 0.27 SD in 2001; 1.33 m OD

± 0.33 SD in 2000; 1.07 m OD ± 0.22 SD in 1999; and 1.17 m OD ± 0.18 SD in

1998. The large standard deviations are a result of the higher surface elevations of

the old saltmarsh which remains along the base of the sea wall. From this it can be

seen that the average elevation of the inner shore has been increasing steadily since

1999 with the largest elevational gain seen between 2000 and 2001, though this will

have been largely due to the sediment top up in November 2000 (Rossi 2001).

The elevation of the inner shore south of grid coordinate 234790 (T1-T5) is higher

than that to the north (T6-T8) by an average of 0.26 m, with a difference in

elevation of 0.42 m between the highest point (T4) and the lowest point (T8). This

pattern in surface elevation was also seen by Rossi (2001) who reported a difference

of 0.50 m between the southern and northern ends of the inner shore. This trend was

not reported in previous surveys although French & Watson (1999) did record lower

elevation levels at the northern and southern ends of the site compared with that at

the middle of the site.

31

-1.50

-1.00

-0.50

0.00

0.50

1.00

1.50

2.00

2.50

3.00

0 20 40 60 80 100 120 140

Distance (m)

Elev

atio

n (m

OD

)

20062001200019991998MHWNMHWS

Figure 12: Change in elevation profiles across Transect 1 (T1) 1998-2006.

-2.00

-1.50

-1.00

-0.50

0.00

0.50

1.00

1.50

2.00

2.50

3.00

0 20 40 60 80 100 120 140

Distance (m)

Elev

atio

n (m

OD

)

20062001200019991998MHWNMHWS

Figure 13: Change in elevation profiles across Transect 2 (T2) 1998-2006.

32

-2.00

-1.50

-1.00

-0.50

0.00

0.50

1.00

1.50

2.00

2.50

3.00

0 20 40 60 80 100 120 140

Distance (m)

Elev

atio

n (m

OD

)

20062001200019991998MHWNMHWS

Figure 14: Change in elevation profiles across Transect 3 (T3) 1998-2006.

-2.00

-1.50

-1.00

-0.50

0.00

0.50

1.00

1.50

2.00

2.50

3.00

0 20 40 60 80 100 120 140

Distance (m)

Elev

atio

n (m

OD

)

20062001200019991998MHWNMHWS

Figure 15: Change in elevation profiles across Transect 4 (T4) 1998-2006.

33

-2.50

-1.50

-0.50

0.50

1.50

2.50

3.50

0 20 40 60 80 100 120 140

Distance (m)

Elev

atio

n (m

OD

)

20062001200019991998MHWNMHWS

Figure 16: Change in elevation profiles across Transect 5 (T5) 1998-2006.

-2.00

-1.50

-1.00

-0.50

0.00

0.50

1.00

1.50

2.00

2.50

3.00

0 20 40 60 80 100 120

Distance (m)

Elev

atio

n (m

OD

)

20062001200019991998MHWNMHWS

Figure 17: Change in elevation profiles across Transect 6 (T6) 1998-2006.

34

-2.00

-1.50

-1.00

-0.50

0.00

0.50

1.00

1.50

2.00

2.50

0 20 40 60 80 100 120

Distance (m)

Elev

atio

n (m

OD

)

20062001200019991998MHWNMHWS

Figure 18: Change in elevation profiles across Transect 7 (T7) 1998-2006.

-1.00

-0.50

0.00

0.50

1.00

1.50

2.00

2.50

0 10 20 30 40 50 60 70 80

Distance (m)

Elev

atio

n (m

OD

)

20062001200019991998MHWNMHWS

Figure 19: Change in elevation profiles across Transect 8 (T8) 1998-2006.

35

5.3 Movement of the gravel bund

Figure 20 shows the outline of the gravel bund on a contour map from the 2006

survey from which the inner edge of the bund was estimated. Although most of the

lateral movement occurred during the first year after the initial recharge between

1998 and 1999, landward migration of the bund is still happening but at a much

slower rate (Figure 21). The fastest rate of migration was seen between 1998 and

1999 where French & Watson (2000) reported the mean landward movement of the

bund to be 13.3m ± 2.8 SD, which slowed to 5.1m ± 2.5 SD between 1999 and

2000. Rossi (2001) reported further migration of the southern part of the bund by up

to 8m between 2000 and 2001 and 2.3m at the northern end, giving an average

migration rate of 6.25 m year-1 between 1999 and 2001. This rate of migration had

slowed even more between 2001 and 2006 where the inner edge of the gravel bund

measured at 2006 was recorded to be on average 7.12 m further inland from its

position in 2001 and the rate of migration was calculated to be 1.42 m year-1. The

migration has not been uniform across the length of the bund and the 2006 survey

shows that the same trend in bund movement as previous surveys with the southern

end of the bund (south of grid coordinate 234790) showing faster rates of migration

than the northern end of the site.

Height of Gravel Bund/ m OD Transect 2006 2001 2000 1999 1998T1 1.94 1.95 1.73 1.52 0.75T2 2.49 2.55 1.84 1.31 1.25T3 2.48 2.18 1.86 1.73 1.14T4 2.36 2.14 2.06 1.53 1.00T5 2.31 2.28 2.12 1.58 1.03T6 2.23 2.12 1.88 1.33 0.86T7 2.17 2.13 1.64 1.25 1.23T8 2.33 2.19 1.68 1.4 1.05Average 2.29 2.19 1.85 1.46 1.04

Table 5: Height of gravel bund on T1-T8, 1998-2006.

36

As a result of this migration, the width of the inner shore has now been reduced to

approximately 28-36 m (average width of 32.5 m ± 2.78 SD) (Figure 20) from the

original 60-70m in 1998 (French & Watson 1999).

There is also a vertical component in the bund movement and all transects indicate

that the height of the gravel bund is now above the MHWS tide. Average heights of

the gravel bund between 1998 and 2006 on each transect are shown in Table 5. The

crest of the bund has increased in height between each survey year with the largest

increases in height occurring in the first 3 years of the survey. The average crest

height of the gravel bund recorded in 2006 was 2.29 m OD ± 0.18 SD which is an

increase of 1.25 m between 1998 and 2006. However, the increase in crest height

between 2001 and 2006 is significantly lower with an average increase of 0.10 m.

5.4 Vegetation survey

Photographs of the site in 2006 (Plates 1 & 2) show that since the last survey in

2001, the inner shore has been colonized quite extensively by saltmarsh vegetation,

namely by Spartina and Salicornia spp. A small dataset of elevation versus

vegetation type recorded patches of bare mud at the lowest elevations of 1.56 m OD

± 0.04 SD, followed by Spartina at 1.59 m OD ± 0.02 SD, and then Salicornia at

1.73 m OD ± 0.15 SD. The elevations of the areas supporting saltmarsh vegetation

are well above MWHN tide which suggests that factors other than elevation may be

important in the establishment of vegetation at the site. Again there are differences

seen between the southern and northern parts of the inner shore. While the width of

southern part of the site was quite well vegetated, this colonization did not extend

over the entire length of the site and the northern end was still devoid of vegetation

(Plates 2 & 3).

37

Figure 20: Inner edge of gravel bund 2006 drawn against elevation, contours and data points.

38

Figure 21: Migration of inner edge of gravel bund between 1998 and 2006.

39

Plate 1: Inner shore looking north 2001 survey (between T2 and T3). Photograph shows remnants of old

saltmarsh and associated vegetation along the base of the sea wall but little vegetation on the mud surface of

the inner shore.

Plate 2: Inner shore looking north 2006 survey. Photograph shows the width of the inner shore has been

colonized extensively by vegetation.

40

Plate 3: Photograph of the northern end of the inner shore in 2006 which is still devoid of vegetation.

41

5.5 Summary of results

The analysis of the results can be summarized as follows:

• The inner shore of the recharge site has experience a net gain in elevation of

0.05 m between 2001 and 2006 and a total increase of 0.62 m since 1998;

• The average elevation of the inner shore in 2006 is now 0.67 m above

MHWN tide at 1.82 m OD ± 0.19 SD compared with 1.78 m OD ± 0.27 SD

in 2001; 1.33 m OD ± 0.33 SD in 2000; 1.07 m OD ± 0.22 SD in 1999; and

1.17 m OD ± 0.18 SD in 1998.

• The elevation range of the outer shore decreased by 0.21-0.37 m between

2001 and 2006 while comparison of the area between 1998 and 2006

revealed an elevational gain at its inner edge (closest to the bund) of 1.03 m

and a elevational loss at its outer edge (closest to the mean low water mark)

of 0.43 m. The increase in elevation is entirely due to the landward

migration of the gravel bund exposing the elevated inner shore to the

seaward edge;

• The gravel bund has migrated a further 7.12 m inland between 2001 and

2006 reducing the width of the inner shore to 28-36 m from the original 60-

70 m in 1998.

• The crest height of the bund increased by 0.10 m between 2001 and 2006

and is now at a height above MHWS and has effectively created a back

barrier marsh system which is protected from wave action and open water.

42

• Vegetation cover across the width of the site is quite extensive with the

dominant species being Spartina at average elevations of 1.59 m OD ± 0.02

SD and Salicornia at average elevations of 1.73 m OD ± 0.15 SD.

43

6 DISCUSSION

6.1 Historical marsh loss and possible causes

In order to successfully manage and restore intertidal habitats we must first understand

the local causes of erosion. Saltmarsh erosion is often associated with sea level rise

(French & Moller 2001) and given the current picture of climate change it is not

surprising that there has been a proliferation of research on this topic in recent years. As

already mentioned in section 2, the natural response of saltmarshes to rising sea levels

is to maintain their relative position in the tidal frame through vertical accretion and

landward migration provided that there is an adequate sediment supply (van der Wal &

Pye 2004). For some areas of saltmarsh in the south-east of England, rates of vertical

accretion are in the order of 2-3 mm year-1 (Wolters et al. 2005) which would mean that

it is keeping pace with the estimated rate of sea level rise in the region of 2-3 mm year-1

(French & Burningham 2003). Unpublished data from CERU recorded sedimentation

rates of 6 mm year-1 in the Orwell estuary which indicates that sedimentation is easily

outpacing sea level rise (French & Burningham 2003) and Guthrie & Cottle (2002)

reported that overall, the Orwell estuary could be seen as an accreting estuary even

though the saltmarshes in the lower estuary are experiencing net erosion.

Saltmarsh accretion is not solely dependent on sediment and land availability and it has

been recognized that saltmarsh development is a complex dynamic process that can be

governed by other factors such as: changes in tidal regime; sediment budgets;

vegetation cover; wind/wave energy; bioturbation; and human activities (Pye 2000,

Moller & Spencer 2002, Hughes & Paramor 2004, Wolters et al. 2005) and a change in

any one of these can also lead to habitat erosion.

The type of saltmarsh erosion observed in the Orwell estuary is mainly a reduction in

the width of the habitat and possible causes associated with this kind of erosion in

estuaries are: dredging activities; increase in wave energy; boat wash; and a change in

the shape of the tidal curve (Carpenter 1995).

44

A literature review on the historical loss of saltmarsh in the Orwell estuary has shown

that most of the loss has occurred in recent years at an accelerated rate. Burd (1992)

analysed erosion and vegetation change of saltmarshes in Essex and north Kent between

1973 and 1988 and showed that 40% of the 100 ha of saltmarsh recorded in the Orwell

estuary in 1973 had been lost by 1988. Of this, 7% was lost through reclamations in the

Felixstowe Docks and Levington marina and the remaining 33% through erosion. Of

the area lost through erosion, 74% occurred in the pioneer zone, while much less was

eroded from the low and low-mid zones, 29% and 22% respectively.

More specifically to the recharge site, Rossi (2001) conducted an analysis of saltmarsh

evolution between 1881 and 1988 by comparing the saltmarsh fringe at Shotley Point

from historical maps imported into ArcView 3.1. The results showed that the seaward

edge of the saltmarsh had on average eroded by 13.4 m during this period, equating to

approximately 0.13 m year-1. However, this landward retreat did not occur at a stable

rate over the years and the analysis indicates that the estuarine system was relatively

stable between 1881 and 1956 with most of the erosion occurring between 1956 and

1988. The results also indicated that erosion of the marsh area appeared to increase with

decreasing distance from Shotley Marina with the parts of the saltmarsh edge retreating

by as much as 41.5 m between 1881 and 1988.

These patterns of erosion seem to coincide with the increase in dredging activity in

Harwich harbour from the 1960’s (Dearnaley 2004). In 1969, HHA required the annual

removal of 50x103 m3 of dredged material from the harbour which increased to 15x104

m3 in 1972 following further developments of Felixstowe in 1971 (Dearnaley 2004).

This figure had doubled by 1981 and reached 2.6x106 m3 in 1995 after the development

of the Trinity I, II and III Terminals (Dearnaley 2004).

Although dredging of estuaries is important in terms of maintaining safe channel

navigation it can have significant impacts on intertidal areas and sets in motion a series

of environmental modifications as the intertidal habitats try to adjust to a new

equilibrium (French 1997, 70). Channel deepening can have similar effects to

45

embanking in estuaries as it increases the tidal range and thus the tidal prism. The tidal

prism is defined as the volume of water exchanged over a single tidal period (Burd

1995) and an increase in tidal prism not only results in decreased bed friction, but also

higher tidal current velocities and thus erosive potential of the tidal flows through the

estuary (Burd 1995, Pethick 1993) with most erosion occurring on the mudflats. As

mudflats generally slope towards the deepest part of the channel, channel deepening

also causes the mudflat profile to steepen through sediment erosion as it tries to adapt to

the new channel depth (Pye & French 1993). Beardall et al (1988) approximated that

39% of the dredged material from Harwich harbour originates from mudflat erosion. A

second impact of dredging is the permanent removal of sediment from the estuary

system which will produce deficits in the sediment budget. As already mentioned, an

abundantly available sediment supply is critical in the survival of saltmarshes against

rising sea levels (Adam 2002). If dredging and subsequent sediment deficit is the

primary cause of saltmarsh erosion in the Orwell estuary, then in theory the practice of

foreshore recharge could be successfully employed as a technique for saltmarsh

restoration and management in the area.

However, the removal of sediment from the Orwell estuary does not appear to be the

main causative factor in the erosion of its saltmarsh habitats. The principle supply of

sediment to the Orwell Estuary is from offshore and passes through Harwich harbour on

the flood tide (Guthrie & Cottle 2002). Bathymetric surveys of the estuary between

1994 and 1999 revealed that the estuary experienced an average vertical accretion of

12-14 mm year-1 upstream of Levington Creek and overall the estuary is accreting at a

rate of 2x104 to 3x104 m3 year-1 (Guthrie & Cottle 2002). Most of the saltmarsh erosion

discussed in this section has occurred in the outer estuary along the open stretches of

Trimley Marshes and Shotley Point (Guthrie & Cottle 2002). This would suggest that

other factors are playing a role in the erosional regime of the outer estuary such as wave

climate.

Very little research has focused on the potential impacts of vessel movements and ship

wash on marine habitats but some common effects have been identified as intertidal

46

erosion within estuaries and sediment re-suspension (ABP Research 1999). Nanson et

al (1994) conducted a study which measured river bank erosion caused by boat

generated waves on the Gordon River in Tasmania and found a direct correlation

between vessel speed and shoreline erosion with erosion observed when wave heights

exceeded 0.35 m. The study also showed that reducing vessel speeds and frequency

decreased the amount of erosion along the river bank. French & Watson (1999) reported

wave heights in excess of 0.25 m occurring regularly along the outer shore following

the passing of commercial ships using the channel and limited field observations

suggested that the largest waves were generated by fast pilot boats and not cargo

vessels. The high speed ferry service between Harwich and the Hook of Holland is also

known to create wave problems along the shores and shallow sandbanks of the

Stour/Orwell estuary which has caused ship wash problems at critical depths and

speeds, although the ferry operator is reported to have taken appropriate action to

address this issue (ABP Research 1999). It is useful to note that the rate of erosion

between 1988 and 1997 was almost half that reported between 1973 and 1988 (Guthrie

& Cottle 2002). Lack of information on dates of imposed speed restrictions on ships

using the harbour means that is not possible to directly correlate this decrease in rate of

saltmarsh erosion to waves created by ship wash.

Port developments such as channel widening can have much the same effect as channel

deepening and results in an increase in wave velocity entering the estuary. Dearnaley

(2004) investigated wave conditions in Harwich harbour using a state-of-the-art SWAN

wave model to examine the effects of channel entrance enlargement on offshore wave

propagation into the harbour following port developments in 1976, 1986, and 2001. A

comparison of the estuary bathymetrics for these years demonstrated wave conditions at

Shotley Point have increased by up to 40% as a result of these changes. In light of the

above, it is plausible to propose that the loss of saltmarsh through erosion seen in the

Orwell estuary between 1973 and 1988 was principally related to the increase in wave

energy caused by the increase in ship activity after the development of the port. This

has been manifested as erosion of saltmarshes in the lower estuary as wave energy is

highest as it penetrates the estuary and attenuates up river.

47

Increased storminess is another factor that needs to be considered when trying to

understand the causes of saltmarsh erosion and a study by Pye (2000) found that recent

losses of saltmarsh in the south-east of England have been caused primarily by an

increase in wind and wave energy. The accelerated loss of saltmarsh in the Orwell

estuary could be related to the findings from this study which reported sustained strong

winds and waves during the periods 1970-1973, 1976-1979, and 1985-1988. Storm

surges entering estuarine areas increases average tidal current velocities and bed shear

stresses which increase the potential for erosion (French 1997, 91). The increase in

wave energy during a particular storm could also initiate damage to the seaward edge of

the saltmarsh leaving it vulnerable to ‘normal’ wave action (Adam 2002).

An alternative explanation for saltmarsh loss in the Orwell estuary is the loss of large

beds of eel grass which were destroyed in the 1930s by a wasting disease (Beardall et

al. 1988). Vegetative die-back can contribute to erosion by de-stabilising the sediment

and exposing the saltmarsh to increase wave attacks. Generally, saltmarsh sediments are

protected from erosion by their vegetation and loss of vegetation can initiate local

erosion (Adam 2002). The presence of saltmarsh vegetation reduces erosion in two

ways: the above ground vegetation traps sediment and encourages accretion as water

flows through at high tide; and the below ground root system binds the sediment

together making it less susceptible to erosion (French 1997, 94). Beardall et al (1998)

suggested that the loss of eel grass beds in Suffolk between 1925 and 1965 could have

played a role in the erosion of the intertidal habitats in the Stour estuary and by the

same reasoning, this hypothesis could also be applied to the erosion experienced in the

Orwell estuary. However, the relationship between eel grass and saltmarsh development

is not definitively clear and most researchers consider eel grass (Zostera spp) to be

functionally and biologically independent of saltmarshes (ABP Research 1997, NRA

1995).

48

6.2 Morphodynamic behaviour of the recharge site

The failure of a number of wetland restoration and creation projects in the US have

been attributed to inappropriate elevation levels of the restored/created site (ABP

Research 1998). In the UK, MAFF (now DEFRA, 1996) suggested that land elevations

most suitable for saltmarsh restoration projects lie between MHWN and MHWS tides

which equates to approximately 450 and 500 tidal inundations per year. Sites created at

elevations lower than MHWN tides will either suffer erosion due to wave action or

result in mudflat creation.

The first application of dredged material onto the recharge site achieved an initial

elevation of 0.9-1.3 m OD in 1998, which was below the MHWN mark of 1.4 m

(French & Watson 1999). The decrease in elevation seen across most of the recharge

site between 1998 and 1999 was to be expected. Fine grained sediments do not develop

an equilibrium immediately but need time to dewater and consolidate before full

cohesive strength and stable elevation is achieved (French 1997). The rate at which

recharge material settles is dependent on the sediment characteristics and the

hydrodynamics of the site.

After the initial elevation loss between 1998 and 1999, the inner shore of the recharge

site started to accrete as subsequent surveys indicated a net gain in elevation. There was

a large gain in elevation seen between 2000 and 2001, although much of this would

have been the result of a top-up recharge in November 2000, which was estimated to

account for 0.40 m of the increase (Rossi 2001). The average gain in elevation between

2001 and 2006 of 0.05 m is equivalent to an annual accretion rate of 10 mm year-1,

which is higher than the expected annual accretion rate of 4 mm year-1 for saltmarshes

in this locality (French 1999). Some of this increase in elevation can be accounted for

by the second top-up recharge between 2001 and 2006 (French J., personal

communication 2006). However, as there is no information on when and how much

dredged material was deposited on the site between these survey years, it is not possible

to report with any accuracy the annual rate of natural accretion of the inner shore and it

49

can only be assumed that the site does accrete naturally based on the increase in

elevation seen between 1999 and 2000 when there was no top-up recharge.

In natural environments, saltmarshes are usually fronted by mudflats which are essential

in the development and stability of saltmarshes and act by dissipating wind, wave and

tidal current energy, thereby providing a low energy environment in which marsh

biological and sedimentary processes can take place (Pethick 1993). This functional

relationship between mudflat and saltmarsh means that any erosion of the mudflat will

ultimately have a negative impact on the saltmarsh. In the case of Shotley Point, the

inner shore is fronted by a gravel bund which protects the area from open water and has

in effect created a back barrier environment which is functionally disconnected from the

outer shore, therefore erosion of the outer shore has not resulted in erosion of the inner

shore. Whilst the gravel bund is proving to be an excellent protective barrier against

wave energy for the inner shore, the long term stability of the site is being threatened by

its continued landward migration. The barrier was a necessity in the design of the

recharge project and was put in place to retain the mud in the recharge site and

minimize sediment re-suspension. However, the mobility of the bund is higher than

what was originally anticipated and is exhibiting a ‘rollover’ process which has seen it

migrate inland by approximately 32.9 m between 1998 and 2006. The movement of the

gravel bund can be compared with that of gravel barriers where the predominant

response to sea level rise is also one of landward retreat (Orford et al. 1995). Wave

energy acting on the barrier can move shoreface material both up to the crest and

vertically incrementing the crest position or over the barrier crest and down the other

side, thus generating bund rollover (Orford et al. 1995). Figure 20 illustrates the lateral

movement of the bund as a result of wave energy where a greater rate of migration is

seen at the southern end of the site, closer to the mouth of the estuary, and is most likely

a result of higher wave energy penetrating the estuary. Even though the crest of the

bund appears to have stabilized and inland migration has slowed to 1.42 m year-1

between 2001 and 2006, the rollover of the crest onto the inner shore has reduced the

width of the shore and is effectively eroding the site in much the same way as coastal

squeeze. The inner shore is now being squeezed between the sea wall and the gravel

50

bund as the rollover process gradually smothers the newly developed habitat. If the

gravel bund continues to migrate at the rate mentioned above, this will result in the

bund rolling over the recharge site and reaching the sea wall in 20-25 years. Not only

does this action question the long term success of the scheme, but the rollover process

will eventually expose the inner shore sediments to the outer shore. Although the newly

exposed sediments exhibit higher shear strength values due to compaction underneath

the gravel bund making them more resistant to erosion, they will eventually be trickle

charged back into the estuary exasperating the need for dredging of the channel which

will in turn lead to further erosion of the outer shore.

From this, it follows that the outer shore does not show the same sediment stability or

accretion rates as the inner shore but has exhibited a net loss in elevation since the first

survey in 1998 with the lowest points in the elevation range deceasing by a further 0.43

m in 2006. The outer shore of the site is effectively a mudflat with an elevation range of

(-1.44)-1.28 m OD and most of the site is covered by MHWN tides. Much of the loss in

elevation has been due to the gravel bund migrating inland and exposing more of the

site to the outer shore. As the bund offers no protection to the outer shore, the continued

erosion of the area is not surprising and is occurring in two ways. Firstly, the outer

shore is still exposed to the wave climate of the estuarine system which was established

to be the main probable cause of erosion in section 6.2. Secondly, in a natural system

where saltmarshes and mudflats are part of the same geomorphic unit, saltmarshes can

provide sediment to be deposited on the mudflat during high wave energy events. This

sediment feedback loop is obviously impeded by the presence of the bund causing the

outer shore to suffer enhanced erosion after storm events with no additional sediment to

replace the loss in elevation. In addition to this, the negative impact of dredging and

channel deepening in the Orwell estuary is likely to continue to steepen the shore

profile and other methods of recharging the outer shore need to be considered.

51

6.3 Vegetation colonization

Vegetation colonization of the mud surface was slow and patchy between 1998 and

2000 and this had not significantly improved by 2001 even though the elevation of the

site lay between MHWN and MHWS tides. However by 2006 the inner shore had been

colonized extensively by halophytic vegetation and the average surface elevation was

calculated to be 0.67 m above MHWN tides at 1.82 m OD ± 0.19 SD. Since there was a

period of 5 years between the 2001 and 2006 survey it is not possible to say when

vegetation colonization began. From field observations it is possible to estimate the

time period required for plant establishment to be 6-8 years after the initial recharge.

Previous surveys by French & Watson (1998, 1999 & 2000) reported a few isolated

specimens of Salicornia spp in the summer of 1998 and 1999 with 20 individual

Spartina anglica plants established in November 1998. This number was reduced in

1999 but the survey in 2000 showed some plants still remained. Rossi (2001) noted that

although saltmarsh vegetation was present, the area it covered was mainly along the

base of the sea wall where there were remnants of old saltmarsh with pre-existing

vegetation, hardly any vegetation was recorded on the mud surface of the site. Although

French & Watson (2000) cited low elevation as being the key factor for the low plant

colonization rate between 1998 and 2000, the elevation of the inner shore in 2001 was

above MHWN tide and at a height suitable for vegetation colonization. This lack of

vegetation could have been due to the instability of the mud surface. Pye & French

(1993) observed that pioneer marsh colonization of a new marsh not only depends on

surface elevation but it also requires a sufficiently stable sediment surface to enable

successful seed establishment and seedling germination. This idea is supported by

Pontee (2003) and a number of methods have been suggested to increase the strength of

dredged material prior to placement which includes mechanical/chemical dewatering

and the use of coarser material which will assist in situ dewatering.

The presence of vegetation on the inner shore will aid the natural accretion process and

this coupled with the top-up recharge between 2001 and 2006 is most certainly the

52

reason for the higher than average accretion rate seen of 10 mm year-1. The difference

in elevation observed between the southern and northern ends of the inner shore can be

expected to increase due to the lack of vegetation colonizing the northern part of the

site.

Elevation Range/m Zone Dominant Species 4 m Tidal Range 3.4 m Tidal Range

Pioneer Spartina anglica 2.30-2.50 1.70-1.90 Salicornia herbacea 2.50-2.90 1.90-2.30 Aster tripolium 2.90-3.35 2.30-2.75 Lower mid-marsh Puccinella maritime 3.35-3.75 2.75-3.15

Halimione portucaloudes

High marsh Limonium vulgare 3.75-4.00 3.15-3.4 Armena maritime Juncus spp

Table 6: Major saltmarsh vegetation zones found in East Anglia and their vertical ranges adjusted to a

standard 4 m tidal range (Boorman 1992) and a 3.4 m tidal range for the Orwell Estuary.

Table 6 above outlines the zonation of saltmarsh vegetation in East Anglia which can be

regarded as representative of south-east England (Boorman 1992) and due the lack of

information on the zonation of Suffolk marshes these figures have been adjusted to the

tidal range of the Orwell estuary against a MHWS tide of 1.85 m OD and a MLWS tide

a -1.55 m OD (Admiralty Tide Tables 2006). From this, it is possible to infer that the

elevation range for pioneer saltmarsh in the Orwell Estuary is 0.15-1.20 m OD. The

elevation at which Spartina and Salicornia spp were recorded in the 2006 survey were

1.59 m OD ± 0.02 SD and 1.73 ± 0.15 SD respectively, which places the vegetation at

an elevation which should be characterized by high marsh vegetation. Normally the

invasion of pioneer species into areas of mid-high marsh zones is a sign of increased

flooding frequency which suggests that vertical accretion is lagging behind sea level

rise (Johnson 2000) however, this is not the case here as the elevation of the site is well

above MHWN tide.

53

Similar findings have been reported at several managed retreat sites where plant

succession in a newly developed area, regardless of the elevation, have followed the

natural sequence usually seen when mudflats initially achieve an elevation that is

suitable for plant colonization (Boorman 1999). In particular, observations from the

large scale managed realignment scheme at Tollesbury in Essex showed that even at

elevations equivalent to that of high marsh the process of re-vegetation followed the

course of succession normally found in the lower marsh (Boorman et al. 1997). The

reason for this at managed realigned sites is thought to be linked to the soil conditions

of the reclaimed land which have changed considerably from those that existed when

the area was originally saltmarsh.

Although this not a suitable explanation for the vegetation succession seen at Shotley

Point, edaphic factors such as organic content and nutrient status of soils are important

in the development and zonation of marsh plants (Gray & Bunce 1972) and can cause

considerable variation in vegetation type in areas of similar submergence. Several

studies have found higher organic matter levels in natural marshes compared with

created marshes and indicate that it can take between 5-25 years for created marshes to

achieve organic content similar to those of natural marshes (Edwards & Proffitt 2003,

Web & Newling 1985). Another factor that may be related to species composition in

saltmarshes is the development of soil micro-flora and the associated micro-fauna

which play an important role in soil processes (Burke et al. 2002). It is possible that a

certain period of time is required for the site to develop the soil conditions required for

the development of high marsh vegetation. Much more research is required on the

subject of vegetation succession on recharged marshes and the factors that cause

deviations between expected and observed species composition at sites such as Shotley

Point before the variation can be fully explained.

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6.4 Success of the scheme

Defining the success of a scheme such as this trial foreshore recharge using beneficial

dredged material is a difficult and controversial issue as the term success is extremely

objective and what may be considered a success by one organization may be deemed a

failure by another. For the purposes of this report, success will be measured at the

simplest level in relation to the original aims of the scheme which were to evaluate the

potential of using muddy dredged material to recharge intertidal areas in the Orwell

estuary measurable in terms of ecological enhancement and flood defence value

(French & Watson 1999).

A flaw in the measurement of restoration success is the limited amount of time the

habitat is allowed to develop before judgment on its success is passed. The timescale

over which success criteria are to be evaluated need to be determined at the outset

(Leggett et al. 2004) and the length of time required to restore a habitat to a steady state

depends heavily on how close the initial conditions of the habitat are to that steady state

(Mitsch & Wilson 1996). Clearly, a heavily eroded habitat will take longer to restore

than a partially eroded habitat. The usual timeframe used to measure restoration success

is approximately 5 years (Mitsch & Wilson 1996). However, due to the stochastic

nature of hydrologic events and slow development of ecosystems, a period of 15-20

years would provide a more accurate picture of how successful a scheme has been

(Mitsch& Wilson 1996). This is certainly true in the case of this trial scheme. HHA

officially ceased to monitor the site in 2000 at which time the site had failed to develop

into a sustainable saltmarsh habitat and remained to be a mud surface with little

conservation or defence value. Although the trial had succeeded in raising the elevation

of the site, based on the results of the 3 year monitoring programme, it could not have

been said with confidence that it had been successful in achieving its original aims. It

has become clear from the last survey in 2006 that a much longer time frame was

required for the site to mature into a relatively stable saltmarsh habitat that is capable of

supporting pioneer vegetation. The inner shore of the site is now one which is providing

a 32 m width of saltmarsh protection to the sea wall while the presence of a pair of

55

Oyster Catchers nesting on the site shows that the recharge has resulted in a certain

level of ecological enhancement and there is potential for the site to develop into a

valuable bird habitat. Since there was a period of five years between the last two

surveys, it is not possible to pin point how long this process took. The extent of

halophytic colonization and the high cohesive strength of the surface observed in the

2006 survey indicates the inner shore has been stable for a number of years. So it could

be suggested that any future trial schemes be monitored for a period of 6-10 years

before a decision on its effectiveness as a coastal defence option is made. It needs to be

said that although the trial has succeeded in enhancing the habitat, the actual

conservation value of the site is debatable, which will be discussed further later on. For

the purposes of establishing a monitoring timeframe to determine the success of the

scheme in terms of ecological enhancement, based on the time sites require to develop

natural processes, it could be suggested that the trial at Shotley should be monitored for

a further 5 years, bringing the total monitoring period to around 15 years.

As mentioned before, the success of the scheme is being compromised by the ongoing

landward migration of the gravel bund, although this has slowed considerably compared

to the first 4 years of monitoring. The slower rate of migration is probably due to a

combination of the recharge mud consolidating into a composite with much higher

shear strength than before and also stabilization of the bund itself in width and height.

As mentioned before, if the bund continues to migrate at the rate seen between 2001

and 2006, it will over a period of 20-25 years displace the recharged habitat, which

would render the scheme unsuccessful. Although a more likely scenario as suggested by

Rossi (2001) could see the dynamics of the bund and inner shore reaching an

equilibrium and forming a self-sustaining marsh/gravel habitat, it should be noted that

patches of shingle vegetation were evident on the gravel bund in 2006. This natural,

albeit unplanned, development of habitats in response to local processes has already

been demonstrated by an intertidal recharge scheme at Pewet Island located in the

Blackwater estuary which used non-cohesive dredged material to recharge intertidal

flats. The post monitoring results showed that lateral erosion of the saltmarsh edge has

stopped in areas that were protected by the gravel bank and the bank itself has provided

56

nesting and roosting sites for birds (ABP Research 1998). Although the gravel bund at

Shotley Point was initially placed there for a purely functional reason, the development

of a habitat that differs from the one planned but which achieves the objectives and

could be naturally sustainable can be viewed as a positive outcome of the scheme.

Although from a conservation value point of the view, the site is less valuable as gravel

supports less invertebrate biomass than mudflats and the continued erosion of the outer

shore is not only further reducing invertebrate biodiversity but will also have a

detrimental effect on feeding sites and feeding times for bird populations.

Despite the possibility that the marsh/gravel habitat may still be a positive development,

future intertidal recharge schemes should investigate methods that would minimize or

prevent bund rollover. One way to achieve this would be to build the bund at a height

and width that is sufficient to prevent wave overtopping. Another possible improvement

on the design of the bund would be to use non-cohesive sediment of a larger particle

size i.e. cobble, which would migrate at a slower rate than the current substrate given

the same wave environment. Osborne (2005) compared gravel and cobble movement at

Grays Harbor, Washington, and reported that at a point between 54 mm and 70 mm size

particles, the cross shore transport rate decreases with increasing size. However, the

study also indicated that larger particles are preferentially moved along shore whilst

smaller particles move cross shore. In light of this, the use of larger cobbles to create

protective bunds may not be appropriate. While they may be slower to move landwards

causing habitat loss through smothering, there is a risk that along shore movement of

the bund would expose the site to local wave energy and thus erosion would ensue.

However, as mentioned earlier, with the current observations, it is not possible to judge

whether the site has successfully created a habitat that is equal in value when compared

to natural marshes in terms of ecological function. The vegetation that has colonized

the site display low species diversity and the continued erosion of the outer shore

remains to be a matter of concern. Mudflats are productive intertidal habitats and are

often neglected in restoration plans, often being incorrectly assessed as habitats that can

suffer erosion and reduced function without contributing to a decline in estuarine health

57

(Short et al. 2000). Although it has been established that the inner shore is protected by

the gravel bund and erosion of the outer shore will not affect the stability of the inner

shore, the conservation value of the outer shore will be diminished with continued

erosion. From the elevation profiles shown in Figure 12-19, the shape of the outer shore

fits the low and concave shape of mudflats described by Kirby (2000) which are

undesirable for the reasons described below. The progressive decline in elevation of the

outer shore will reduce the area that lies between MLWN and MHWS tides classified as

‘middle tidal flat’ by Dryer et al (2000) and the ‘upper tidal flats’ which lie between

MHWN and MHWS tides. The mid-upper tidal flat is generally where the invertebrate

biomass is concentrated and provide important feeding grounds for birds. Continued

erosion and steepening of the outer shore will not only mean a small rise in sea level

will result in a large loss of this important mid-upper tidal flat area, but also reduces

bird feeding time as the low water mark encroaches landwards. Whilst alternative

methods of recharging or protecting the outer shore must be considered if the

conservation value is to be improved, it could be that due to mobility of the sediments

that form mudflats in conjunction with local forcing and constraints i.e. changing sea

levels and dredging, achieving an equilibrium whereby there is no net loss in elevation

may not actually be possible.

In light of the above, more extensive monitoring is required to investigate the long term

stability of the site and also to explore the extent to which ecological functioning of the

site has been restored before it can be said with any level of certainty that this approach

in recharging eroding habitats could be used on a larger scale. Even if further

monitoring proved that intertidal recharge is an effective and viable method of restoring

habitats, the applicability of this technique would be restricted to areas where there has

been a reduction in saltmarsh width due to the same forcing factors discussed in section

6.2.

58

6.5 Limitations of the study

The main limitations to this study were the differences in the datasets collected and are

summarized as follows:

• The number of data points recorded along transects T1-T8 in 2006 were low

compared to the number of measurements recorded in pervious surveys. There was

also a mismatch between the locations of the transects and their locations in

previous surveys, which to a certain extent could not be helped as it was not

possible to locate the exact position of the previous transects. Also, the surface of

the site past T8 had not consolidated enough to enable a safe walkover survey. As a

result of this, much of the data analysed in the results are interpolated from existing

data and there exists a margin of error although general trends were still evident.

• A larger vegetation dataset should have been collected to detail the vegetation cover

and species elevation. The small dataset recorded for this was not extensive enough

to enable any in depth analysis or discussion.

• No invertebrate data were collected and no shear strength data were included in this

study.

In general, more extensive data collection would have provided a better comparison of

the site between the years and a better evaluation of the success of the scheme.

59

7 CONCLUSION

7.1 Summary of findings

• The loss of saltmarsh and recent accelerated erosion in the Orwell estuary is most

likely due to dredging activities, port development, increased storminess and

subsequent increase in wave energy;

• The scheme has been successful in raising the elevation of the inner shore to an

elevation of 1.82 m OD ± 0.19 SD and is now at a level that is sustaining pioneer

marsh vegetation, although the outer shore is continuing to erode;

• The estimated time period required for saltmarsh plant establishment at the site is 6-

8 years;

• The site is now providing a 32 m width of saltmarsh as protection to the sea walland

the presence of one pair of Oyster Catchers nesting on the site indicates the potential

of the site to develop into a valuable bird habitat;

• The gravel bund is continuing to migrate inland and could either (1) displace the

newly evolved inner shore, rendering the scheme unsuccessful; or (2) stabilize and

form a self sustaining marsh/gravel habitat; future schemes should investigate the

use of heavier substrates or improve the design of the bund to minimize movement;

• Although the site has been colonized by vegetation typical of the pioneer zone, the

elevations at which they were recorded are usually characterized by high marsh

species which could be attributable to edaphic conditions such as organic content,

nutrient levels and the development of micro-flora and associated micro-fauna;

• Created saltmarshes can take between 5-25 years to develop similar organic matter

levels as natural marshes, therefore future intertidal recharge schemes should be

60

monitored for a period of 15-20 years before the success of the scheme can be

evaluated;

• Improvements to the design of future recharge projects include: chemical

dewatering of recharge sediment before placement improved design of gravel bunds

to minimize cross shore migration.

In conclusion, intertidal recharge is not a rapid process and it can sometimes take years

for the site to settle at an elevation that is suitable for saltmarsh formation. In terms of

aiding managed realignment strategies this technique could potentially raise surface

elevation to levels that would support saltmarsh vegetation, although there is a

requirement for further research into saltmarsh plant colonization and species

composition on restored sites. Further monitoring is required before it can be said with

a higher level of confidence that this method of restoring habitats could be used on a

larger scale at eroding sites within estuaries or of a similar nature to those in the Orwell

estuary.

7.2 Recommendations for future research

• Further topographical surveys concentrated in the outer shore to monitor rate of

erosion and investigate restoration options; surveys should also monitor the

horizontal and vertical movement of the gravel bund;

• More extensive flora survey and a comprehensive fauna survey is required to

compare the ecological value of the site with that of a natural marsh;

• Investigate organic content and nutrient status of the soil to determine if these are

factors limiting the vegetation species composition pioneer species;

61

• Investigate wave heights and velocities at the site and their sources to establish the

erosive effect on the outer shore;

• Investigate the possible causes of the differences in elevation experienced between

the northern and southern parts of the site.

62

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