LITERATURE REVIEW OF DREDGING ACTIVITIES: IMPACTS, MONITORING AND MITIGATION

ME1101

Development of Approaches, Tools and Guidelines for the Assessment of the Environmental Impact of Navigational Dredging in Estuaries and Coastal Waters

Literature Review of Dredging Activities: Impacts, Monitoring and Mitigation

Contributed by:

Chris Vivian

Andrew Birchenough

Neville Burt

Stefan Bolam

Dean Foden

Ruth Edwards

Karema Warr

Daniel Bastreri

Lara Howe

Contents

Introduction and project outline1

Objective 2: Literature review of significant impacts of dredging activity

and it’s monitoring2

1. Dredging, dredger types and associated impacts2

2. Release of contaminants during dredging7

3. Effects of navigation dredging in water quality: nutrients, dissolved

oxygen and re-suspension of bacteria13

4. Radionuclides in dredged material20

5. The impacts and monitoring of maintenance dredging - related to

turbidity22

6. Capital dredging impacts31

7. Mathematical models used to assess the impact of dredging activity34

Objective 3: Literature review of existing mitigation measures and their efficacy in addressing these impacts 39

References 59

Introduction and Project Outline

Navigational dredging is a fundamental and essential activity for most ports and harbours in the UK to maintain navigable depths for vessels. It is vital to social and economic development of a country that is heavily dependant on maritime trade. In the UK 25–50 million tonnes (wet weight) of material are dredged for disposal to licensed sites each year. Navigational dredging can impact both the seabed through disturbance and removal of sediment and the water column through re-suspension of sedimentary material. Areas subject to regular maintenance dredging are likely to have disturbed or impoverished seabed faunas but this would not be the case for areas previously un-dredged or occasionally dredged. The potential environmental impacts of navigational dredging include effects on marine organisms due to increased turbidity in the water column and subsequently effects on benthic organisms when the suspended sediment settles out, biotic effects from the mobilisation of contaminants, organic matter, bacteria/viruses and nutrients from the disturbed or suspended sediment into the dissolved state and potentially associated reductions in dissolved oxygen. In addition, dredging will have seabed morphological effects and may induce changes in sediment transport rates by reducing the stability of seabed sediments and mobilizing sediments via turbidity plumes.

With the advent of the Marine and Coastal Access Act 2009 and the forthcoming Marine Strategy Framework Directive, as well as the implementation of the Water Framework Directive, it is important we can be in a position to carry out sound assessments of the impacts of dredging and understand the available measures that can be used to mitigate those effects in order to ensure effective integrated management of the marine environment.

A lot of work has been done in assessing dredging impacts, particularly in the US. Therefore the first step was to carry out a literature review of significant impacts of dredging activity and it’s monitoring and of existing mitigation measures and their efficacy in addressing these impacts to assess what R&D issues we need to address before making specific proposals. We later propose to develop approaches, tools and associated best practice guidance so that we can effectively assess the environmental impact of proposals for navigational dredging to contribute to the integrated management of the marine environment.

This review document outlines the current information available on impacts, monitoring and existing mitigation and highlights any gaps in the existing knowledge associated with dredging activities.

OBJECTIVE 2: Literature review of significant impacts of dredging activity and it’s monitoring

1. Dredging, dredger types and associated impacts

1.1 Introduction

Dredging can be described as the process of removing part of the seabed or its overlying sediments with the aim of deepening the area commonly for the purposes of navigation or associated with construction projects.

There are two main types of navigational dredging activities; capital dredging and maintenance dredging.

Capital dredging is the removal of material to create a greater depth than had previously existed. The Marine Management Organisation (MMO) definition of capital dredging is:

‘Material arising from the excavation of the seabed, generally for construction or navigational purposes, in an area or down to a level (relative to Ordnance Datum) not previously dredged during the preceding 10 years.’

Maintenance dredging is required to maintain water depths in areas where sedimentation occurs. It mainly involves the removal of recently deposited unconsolidated sediments, such as mud, sand and gravel. It is undertaken by many operators (ports, berth operators and marinas) to maintain navigable channels and berths. It is generally an ongoing activity that can consist of cycle or a series of repeat dredges. The MMO definition of maintenance dredging is:

‘Material (generally of an unconsolidated nature) arising:

· From an area where the level of the seabed to be achieved by the dredging proposed is not lower (relative to Ordnance Datum), than it has been at any time during the preceding 10 years; or

· From an area for which there is evidence that dredging has previously been undertaken to that level (or lower) during that period.’

1.2 Dredging methods

Dredging used in the marine industry for a variety of navigational purposes ranging from routine maintenance of small marinas and harbours to the creation and deepening of navigation channels and berths at major ports. The choice of dredging plant is largely dependant on environmental conditions such as the hardness and quantity of material to be dredged, site exposure, the method of disposal etc.

Different types of dredging equipment and techniques are employed to achieve the required project outcomes in the most efficient way. There are 3 main dredging methods which are based on the physical processes involved in the excavation and transport of the dredged material, these are described below:

Hydraulic dredgers: use a centrifugal pump and pipe system to raise loosened material in suspension to the surface. There are three main types of hydraulic dredger, suction dredgers (SD), cutter suction dredgers (CSD) and trailer suction hopper dredgers (TSHD).

Mechanical dredgers: use of mechanical excavation equipment to loosen the seabed sediment and raise it to the surface. There are three main sub-groups of mechanical dredger, grab dredger (GD), backhoe dredger (BD) and bucket ladder dredger (BLD).

Hydrodynamic dredgers: use the re-suspension of sediments and their transport away from the dredge area by means of natural forces. The term ‘Hydrodynamic dredging’ is often used to group the following dredging techniques, water injection dredgers (WID), forms of agitation dredgers, that use mixing to make a density current and underwater plough dredgers (UPD), which stir or rake sediments into suspension.

The main types of dredgers used for capital and maintenance dredging are briefly described below:

1.3 Dredgers used in capital dredging projects

Capital dredging projects are often large-scale projects which involve the creation or deepening of access channels and/or berths. With capital dredging the full range of materials may be encountered and soft materials, such as clays, sands and silts, but generally the material dredged has been undisturbed for several years and therefore tends to be consolidated such as stiffer clays, boulders and rocks.

Cutter suction dredgers (CSD) are static dredgers usually associated with capital dredging or areas of harder soils, which have to be removed in thick layers (Bray 2008). The cutter suction dredge is usually an accurate method of dredging (Eisma (Ed) 2006), as it is limited to where the head is deployed.

The cutter suction dredge dislodges material with a rotating cutter equipped with cutting teeth. The loosened material is sucked into the suction mouth located in the cutter head by means of centrifugal pump installed on the pontoon or ladder of the dredger. Further transport of the material to the relocation site is achieved by hydraulic transport through a discharge pipeline. Occasionally the material can be pumped into transport barges for further transport (Bray 2008).

Most cutter suction dredgers do not have an optimal combination of cutting capacity and suction capacity for all types of soil, thus contributing to sediment resuspension (Dearnaley et al 1996; Bray 2008). Also a spill layer (0.25 – 1m) remains in general on the seabed after dredging if no special precautions are taken. An additional pass at the same dredging depth can remove this spill layer. The resuspension caused by the cutter suction dredger can, however be reduced by the following considerations:

· Cutter speed, swing velocity and suction discharge must be optimised with respect to each other. The continuous improvements in automation, control and the cutter suction head positioning have afforded considerable economic and environmental advantage.

· A moveable shield around and above a cutter head or suction head reduces the escape of suspended material into the surrounding water column.

· Optimisation of the design of the cutter head with respect to the material being dredged to improve the direction of the material toward the suction intake.

· The use of silt curtains (Dearnaley et al 1996).

Backhoe dredgers (BD) are basically pontoon mounted or shore-based excavators or diggers that are used to dredge marine sediments. Sediment is excavated by the crane bucket which is then raised above water by the crane arm. This method of dredging can be highly accurate which may be of benefit when dredging contaminated sediment or in environmentally sensitive areas. The excavated material is then placed in barges or trucks for transport.

Bucket Ladder dredger (BLD) consists of a large pontoon with a central well in which a ladder equipped with an endless chain of buckets, is mounted. During dredging, the endless chain rotates along the ladder. The lowest bucket digs into the bed material and the cut material falls into the bucket. It is then carried upwards as the bucket chain rotates. This type of dredger is used infrequently in dredging projects today.

1.4 Dredgers used in maintenance dredging projects

As outlined above maintenance dredging mainly involves the removal of recently deposited unconsolidated sediments, such as mud, sand and gravel. The main type of dredger used for maintenance dredging in the UK and Europe is TSHD, estimated to be responsible for in excess of 90% of dredging by volume in Europe (Bates, 2005). Most ports that have to carry out substantial maintenance dredging each year use TSHD, they are mainly used for dredging of unconsolidated sediments of lower strength. Small ports, harbours and marinas tend to use mechanical dredgers such as GD and BD working with barges. These dredgers are used in more confined areas, such as alongside wharfs, because they are smaller than suction dredges. Until recently WID had limited use in the UK but they are being used increasingly to dredge ports and harbours around the UK (Sullivan, 2000).

The main types of dredgers used for maintenance dredging are briefly described below:

Trailing Suction Hopper Dredgers (TSHD) are most commonly used for maintenance dredging in coastal areas. They are normal sea going ships with a large hopper and equipped with single or double trailing suction pipes that end in a draghead. The draghead may incorporate a water jet system, blades or teeth, or other means of dislodging the material. The function of the draghead is to allow the material to flow to the suction inlet as efficiently as possible as the ship moves slowly forward. The material is lifted through the trailing pipes by one or more pumps and discharged into a hopper contained within the hull of the dredger. Horizontal transport is achieved by the ship navigating to the site where the material is to be relocated, the hopper doors are then opened and the sediment descends to the seabed.

Grab dredgers (GD) are relatively simple and involve the collection of sediments in a crane-mounted bucket, the jaws of which are opened and closed to trap the sediment. Depending on the type of material to be dredged different grab bucket designs can be employed, such as mud grabs, sand grabs and heavy digging grabs.

Backhoe dredgers (BD) are described above and can be used for both capital and maintenance dredging projects. BD can result in increased sediment suspension during the raising of the bucket. Spillage can occur throughout the complete height of the water column. This can be limited by reducing the speed of the bucket line. The bucket leaving a clean surface carries the majority of the loosened soil away, however there is a risk of a spill layer remaining if there is excessive spillage whilst the bucket is lifted and the operator must maintain the optimal horizontal position in order to prevent spillage.

Water injection dredgers (WID) inject high quantities of water under low pressure into bed sediments to reduce their density to the point that they act as a fluid and flow over the bed naturally through the action of gravity. The distance the water-sediment mixture travels is dependent on a number of factors including the sediment composition, density, morphology and the gradient of the bottom slope (Van Raalte, and Bray, 1999).

Underwater plough dredgers (UPD) also know as bed levellers have a basic design of a large frame or steel bar that is pulled over the bed. The frame has a cutting blade which scrapes the bed cutting the bottom sediment layers, these build up in front of the frame and are moved away from the dredge area.

1.5 Associated impacts

Navigation dredging operations have the potential to effect the environment in a number of ways ranging from impacts to habitats and species and effects on physical processes to disturbance to humans and birds from noise. However not all these effects will necessarily cause an impact on the marine ecosystem. Whether an effect actually causes an impact depends on a number of factors such as frequency, duration and magnitude.

The effects of the dredging process can be both direct/indirect and short/long term. Direct impacts of dredging can include direct removal of habitat and species and smothering of benthic habitats, indirect impacts include contaminant release through resuspension of sediments and changes to hydrodynamics and sediments regimes. Impacts can be short term, for example an increase in turbidity due to excavation of sediment or the direct removal of a habitat, or long-term such as changes in flow rate.

Details of these impacts can be found in the sections below and also in chapter 6 ‘Machines. Methods and Mitigation’ pages 127-190 in Bray (2008).

The following sections focus on the impacts associated with maintenance dredging however capital dredging can result in the same impacts but can also have additional impacts, therefore these are discussed in section 6.

2. Release of Contaminants during Dredging

2.1 Introduction

Releases of contaminants associated with dredging can occur in particulate, dissolved or volatile fractions, each characterised by differing transport and/or exposure pathways (Thibodeaux 2005). Variation exists in the liberation mechanism exhibited for different classes of pollutants. When sediments contaminated with heavy metals are dredged, their interactions with iron chemistry may temporarily prevent the partial dispersion of their dissolved form. In the longer-term post dredging of such sediments, substantial release of heavy metals into the water column is to be anticipated. In the case of organic micropollutants, a substantial amount of dissolved organic bound matter pollutants may enter the water column during dredging ( Goossens and Zwolsman 1996).

The degree of contamination of sediments clearly plays a very important role in determining the significance of any mobilisation of contaminants from those sediments. This applies to all types of contaminants including those mentioned in this and subsequent sections of the review. In this context it should be noted that the Environment Agency’s ‘Clearing the Waters’ guidance for assessing compliance with the Water Framework Directive utilises in its assessment Cefas Action Levels for metals, TBT, halogenated organics and PAHs in sediments.

2.2 Heavy Metals

The majority of metal contaminants partition onto particulate matter such as clay minerals, Fe and Mn oxides/hydroxides, carbonates, organic substances (e.g. humic acids and biological materials (e.g. algae and bacteria) (Calmano et al 1993). Fe and Mn oxides/hydroxides along with organic matter are important binding sites for metals in oxic sediment (Saulnier and Mucci 2000; Li et al 2000; Zoumis et al 2001; Fan et al 2002) and that the formation of metal sulphides dominates in anoxic sediments (Di Toro et al 1990; Zhuang et al 1994; Caetano et al 2002). Metal mobilisation and immobilisation processes in sediment are controlled by characteristics of the physicochemical environment (pH, redox potential (Eh) and salinity), sediment properties (clay content, organic matter content, quantity and type of cations and anions, amount of potentially reactive iron and manganese) and microbial activity. When dredging disturbs the environment, the processes affecting release mechanisms of metals from the sediment are altered. Where anoxic sediment is oxidised, redox conditions change transforming the nature of the heavy metal bonds. This in some instances liberates new chemical forms of these metals and in other cases immobilisation is promoted. In partially oxidised sediments where Eh and pH do not change radically, the release of metals is negligible (Forstner et al 1989; Reible et al 2002). Metal release is multiphasic with one set of processes controlling early resuspension (first few hours) and different processes controlling long term release over the course of weeks Reible et al 2002). Periodical or continuous cycling processes lead to significant fluctuations in the concentrations of sediments (Latimer et al 1999).

2.3 Organic compounds

Organic and organometallic contaminants preferentially partition to organic matter in sediment and dissolved organic matter (DOM) in pore water (Goossens and Zwolsman 1996).Therefore, compounds such as pesticides, PCB’s and PAH’s are most often adsorbed to particles or organic matter, a physical condition often caused by the very low water solubility and weak polarity of the organic compounds. Many such compounds do not occur naturally in aquatic environments and are almost exclusively of anthropogenic origin. Unlike metals, organic compounds undergo relatively slow natural degradation as a result of bacterial activity of chemical phenomena such as hydrolysis or photolysis. Organic compounds may react differently dependent upon environmental conditions, e.g. evidence exists that PAH’s tend to leach more under anaerobic conditions than under aerobic conditions (Brannon et al 1993).

When entering an aquatic system PAH’s distribute between different phases including truly dissolved, colloids, suspended particulate matter, surface sediments and biota. Their distribution between these phases depends on their intrinsic physicochemical properties including solubility, vapour pressure and lipophilicity. King et al (2004) showed that PAH behaviour is highly complex in a marine system, and controlled by the interplay of PAH sources, compound physicochemistry, water and sediment movement and field conditions. They noted that dissolved concentrations of PAH’s increased after periods of dredging and severe rainfall. They concluded that external factors can have a powerful influence on contaminant behaviour and even compounds bound to sediment and theoretically removed from the water column can be remobilised and reassert their influence on marine pollution.

Generally, hydrophobic organic contaminants readily desorb from the sediment, although the rate of desorption tends to decrease with time (Chen et al 1999; Lamoureux and Brownawell 1999; Zhang et al 2000). Desorption rates and times also depend on the size of the sediment particles and in the case of PCB’s, degree of chlorination (Borglin et al 1996). Historically contaminated sediments may also exhibit slow or highly resistant contaminant desorption (Eggleton and Thomas 2004). High molecular weight PAH’s are thought to be associated with larger particle sizes and will therefore have different fate to low molecular weight PAH’s (Latimer et al 1999).

Desorption of TBT is highly dependant on both pH and salinity, with highest desorption occurring at low and high pH and intermediate salinities (approx 30%). TBT desorption is lowest in freshwater sediments with pH 7 and increases with influx of saline water (Langstone and Pope 1995). Little or no data are available on the release of organometallic compounds from sediments during resuspension (Eggleton and Thomas 2004).

2.4 Physical processes

Fine particles resuspended during the dredging process can remain in the water column for many hours due to their low settling velocities. This material and any associated contaminants can be transported from the dredging area into the surrounding environment due to the effects of currents. Sediment re suspension is also responsible for the release of contaminants into the water column in their dissolved state from pore water and desorption of contaminants from suspended sediment particles. This release is particularly significant given the ready bioavailability of the dissolved contaminant phase (Eggleton and Thomas 2004). The magnitude and temporal extent of the risks depend on a number of factors including duration of the dredging operation, composition of the sediments being dredged (e.g. grain size distribution), contaminants associated with the sediment, current velocities and a range of other physical and chemical factors (Bridges et al 2008).

Other transport mechanisms of contaminants may exist during the dredging process in addition to input to the water column, e.g. release to the air via volatilisation. Floating oils can also be released to the water column during the dredging process e.g. non-aqueous phase liquids.

Understanding of contaminant release is limited with very little empirical data available on the magnitude of contaminant releases and contaminant release mechanisms. Contaminant releases have been estimated from measurements of dissolved and total contaminant concentration from samples collected from a sparse spatial grid with limited frequency (Steuer 2000; Alcoa 2006).

Dissolved contaminants exiting the dredging zone will attempt to establish a concentration in equilibrium with solids or organic carbon in suspension. The mass of these dissolved contaminants in the water column changes slowly due to the time take to reach equilibrium. Contaminant concentrations exhibit great variations in their temporal, vertical and lateral distribution due to variability in the dredging operations and dilution by turbulent diffusion in the water column (Hayes et al 2000).

2.5 Summary of fate and transport processes

In the far field, contaminant fate and transport processes are not unique to dredging operations, the dominant processes are (after Bridges et al 2008): -

· Settling of particle-associated contaminants

· Bed erosion/resuspension, exchange with suspended and bed loads and deposition/burial

· Hydrodynamics

· Partitioning and kinetic rates of adsorption and adsorption/desorption

· Bioturbation

· Molecular diffusion

· Groundwater advection

· Volatilisation

· Biogeochemical transformation (e.g. oxidation, complexation, precipitation, biotic and abiotic transformation, diagenesis etc.)

In the near field processes affecting contaminant concentrations are predominately

· Erosion

· Hydrodynamics (advection and turbulent diffusion)

· Settling (dredging operation dependant due to flocculation and shearing of aggregates at source)

· Partitioning (dredging operation dependant due to desorption kinetics of aggregates)

2.6 Bioavailability of contaminants as a result of dredging activity

Bioavailability with respect to Dredge related bioavailability is mainly site specific and dependant on the degree of contamination, the amount of suspended sediment , the duration of the disturbance and the organism (Su et al 2002) .There are a number of ways that contaminants become available to aquatic organisms; through ingestion with food, through membrane-affiliated transport or through passive diffusion. The rate of uptake and mechanisms for uptake vary among and within species and depend on a number of factors including development stage, sexual condition and history of contaminant exposure.

The degree of contaminant bioavailability is determined by ‘the reactivity of each contaminant with the biological interface, the presence of other chemicals that may antagonise or stipulate uptake, and external factors, such as temperature, that affects the rate of biological or chemical reactions (Louma 1983). Bioavailability is also influenced by the ability of an organism to metabolise contaminants. For example PCB’s tend to be acutely toxic to sediment dwelling organisms, especially when they occur in mixtures with other contaminants (McDonald et al 2000).

Bioavailability and bioaccumulation of contaminants in the aquatic environment is therefore mainly dependant on the partitioning behaviour or binding strength of the contaminant to sediment. Dissolved or weakly adsorbed contaminants are more bioavailable to aquatic biota compared to more structurally complex mineral-bound contaminants (Calmano et al 1993).

The bioavailability of organic contaminants in aqueous exposure pathways decreases with increasing octanol/water partitioning coefficient due to hydrophobic properties of the chemicals involved. The primary route for sediment uptake is then ingestion.

It has been suggested by Gschwend and Hites (1981) and Farrington et al (1983) that oil-associated PAHs might be more available for microbial digestion than PAH’s associated with combustion and that Petroleum based PAH’s were more available for uptake by mussels than pyrogenic PAH’s. The inference being that environmental behaviour and bioavailability are directly source dependant.

The only accurate way of determining metal bioavailability is to perform bioassays, including bioaccumulation tests and/or toxicity identification evaluation (TIE). The limitation is that toxicity is related to the sensitivity of the test animal and the length of the test.

Bocchetti et al (2008) found that the bioavailability of trace metals and PAH’s increased during dredging activities in Piombino (Tuscany, Italy), with values up to 40μg/g for PB and up to 2200ng/g for PAH’s in tissues of caged mussels. Whilst bioavailability of trace metals returned to the pre-dredging values after the cessation of dredging the accumulation of PAH’s, oxidative effects and genotoxic damages remained elevated in mussels caged in the inner harbour area. This suggests that lipophilicity and sedimentation rate of chemicals can influence the duration of impact after their re-mobilization from contaminated sediments.

2.7 Mitigation considerations when dredging contaminated sediments

Dredgers have been specially designed for working in environmentally sensitive sites, such as those where contaminated sediments have been identified. New vessels are continually being developed together with new techniques for environmental mitigation.

At the dredging site itself in addition to the careful planning and execution of the dredging actions, physical barriers can also be employed to prevent the spread of suspended sediments. These ‘barriers’ can be erected at or near the dredging site and often consist of silt screens/curtains which control the dispersion of turbid water by diverting the flow under the curtain, thereby minimising the turbidity in the upper layers of the water column outside the silt curtain. The utilisation of a bubble curtain is sometimes considered as an alternative to a silt screen. The upwelling of bubbles from the sea or riverbed to the surface prevents fine sediments from passing across. Both techniques are only effective where current conditions are relatively slow. Further research is required to identify and describe the extent of mitigation measures available (Bray 2008).

Remedial dredging can be conducted to remove sediments contaminated above certain action levels whilst minimising the spread of contaminants to the surrounding environment during the dredging process. It requires the cautious removal of the contaminated material and is often linked to further treatment, reuse or relocation of these materials. Remedial dredging can only be successful if the source of the contamination is removed prior to this operation. USACE and PIANC, amongst others, have compiled general guidance notes on the application of this approach.

2.8 Summary

Sediment disturbance events, such as dredging and other natural anthropogenic activities can result in alteration in the chemical properties of the sediment. These properties, together with the nature of the pollutant concerned, govern the mobilisation mechanisms for the release of such contaminants.

Contaminant release from dredging operations needs to be quantified to estimate short-term exposures and risk of dredging and potential impacts on long-term risk following completion of dredging. The degree of risk is multi factorial dependent upon methodology and duration of the dredging operation, physical nature of the dredge material, characteristics of contaminants associated with the sediment, current velocities and a range of other physical and chemical factors. Contaminants adsorbed on and to resuspended particles may partition to the water column and be transported great distances downstream in dissolved form along with dissolved contaminants in the released pore water. The resuspended sediment particles will settle and contaminants may be released to the water column by densification, diffusion and bioturbation. This release may have significant implications for the long-term flux of sediment associated contaminants into the water column. Understanding of both contaminant release mechanisms and magnitude of release is limited. Bioavailability and bioaccumulation of contaminants in the aquatic environment are largely governed by partitioning characteristics of the contaminant to sediment, although bioavailability is also influenced by the ability of an organism to metabolise contaminants. The development of new methodology and techniques for mitigation of potentially adverse environmental impacts caused by the dredging of contaminated sediments are ongoing.

The degree of contamination of sediments clearly plays a very important role in determining the significance of any mobilisation of contaminants from those sediments. This applies to all types of contaminants including those mentioned in this and subsequent sections of the review. In this context it should be noted that the Environment Agency’s ‘Clearing the Waters’ guidance for assessing compliance with the Water Framework Directive utilises in its assessment Cefas Action Levels for metals, TBT, halogenated organics and PAHs in sediments.3. Effects of navigation dredging in water quality: nutrients, dissolved oxygen and re-suspension of bacteria

3.1 Organic Enrichment

Human activity has an enormous influence on the global cycling of nutrients, especially on the movement of nutrients to estuaries and other coastal waters. For phosphorus, global fluxes are dominated by the essentially one-way flow of phosphorus carried in eroded materials and wastewater from the land to the oceans, where it is ultimately buried in ocean sediments (Hedley and Sharpley 1998). The size of this flux is currently estimated at 22 Tg P yr−1 (Howarth et al 1995). Prior to increased human agricultural and industrial activity, the flow is estimated to have been around 8 Tg P yr−1 (Howarth et al 1995). Thus, current human activities cause an extra 14 Tg of phosphorus to flow into the ocean sediment sink each year, or approximately the same as the amount of phosphorus fertilizer (16 Tg P) applied to agricultural land each year.

The effect of human activity on the global cycling of nitrogen is equally immense, and furthermore, the rate of change in the pattern of use is much greater (Galloway et al 1995). The single largest global change in the nitrogen cycle comes from increased reliance on synthetic inorganic fertilizers, which accounts for more than half of the human alteration of the nitrogen cycle (Vitousek et al 1997). Use as of 1996 was approximately 83 Tg N yr−1. Approximately half of the inorganic nitrogen fertilizer that was ever used on Earth has been applied during the last 15 years.

Microbially mediated redox reactions in estuarine sediments lead to the remineralization of organic matter and the recycling of nitrogen and phosphorus compounds (Fenchel and Blackburn 1979). In the presence of oxygen, NH4+ released during remineralisation of organic nitrogen compounds is oxidized to NO-2 and NO-3 through the biological process of nitrification. However, in anoxic coastal sediments, NH4+ can accumulate to mM concentrations in sediment porewaters. Estuarine sediments generally are in the pH range of 7 to 8, causing NH4+ to be the dominant form of ammonia. This positively charged ion can participate in adsorption–desorption reactions with sediment solids followed by incorporation into the solid phase absorption. (Rosenfeld 1979; Froelich 1988). This process has been suggested to be an important buffering mechanism for nutrient concentrations in estuaries (Pomeroy et al 1965). Under high pore water NH4+ concentrations, the quantity of NH4+ maintained in the adsorbed phase is primarily limited by the cation exchange capacity of the sediments.

Bottom sediments can supply a significant fraction of the nutrients required by primary producers in estuaries (Nixon 1981; Rizzo 1990). In shallow estuarine systems characterized by frequent resuspension of surface sediments, desorption from suspended particles can also supply NH4+, delaying efforts to reduce the effects of eutrophication (Rizzo and Christian 1996). Release of nutrients from resuspended porewater and sediment particles has been implicated in the stimulation of heterotrophic microplankton in estuarine waters (Wainright 1987). Considering the capacity for elevated porewater nutrient concentrations, this effect is not surprising. Investigations of the extent and the geochemical mechanisms of nutrient release (Blackburn 1997; Rutgers van der Loeff and Boudreau 1997; Wainright and Hopkinson 1997) have mainly focused on organic matter mineralization and mineral dissolution during periodic resuspension of surface sediments.

Dredging activities in coastal regions result in the relocation of large volumes of anaerobic porewater and particles. The method used by the US Army Corps of Engineers (responsible for the maintenance of coastal waterways in the US) to assess the release of NH4+ during dredging involves mixing dredged sediments with overlying water at relatively high solid to solution ratios (1:5 v/v). At these high ratios, the potential NH4+ release may be underestimated.

A series of closed system experiments undertaken by Morin and Morse (1999) resulted in much greater release of NH4+ (g d.wt. -1) when compared to the elutriate method. There are several differences between the two methods which influence the results. First, the elutriate method involves constant aeration resulting in conversion of any easily oxidized compounds (e.g., HS-). Second, according to the elutriate technique agitation of the sediment water mixture is terminated after 1 h whereas sediment suspension is maintained in the closed system. Although these differences in methodolog influenced the results, they are probably of secondary importance to the ratio of sediment added to overlying water outlined by each method. They also found that about two thirds of NH4+released from resuspended sediments is from desorption rather than simple dilution of porewaters. Results were substantially confirmed by direct observation of a dredging event. Although quite approximate, their model calculations raise the possibility that dredging activities in Laguna Madre could locally cause releases of NH4+ to the water column similar in magnitude to those from natural benthic fluxes.

Nutrients dissolved in sedimentary pore water will diffuse slowly to overlying waters. While sediment particles may also bind some nutrients (Nixon et al 1976; Rosenfeld 1979), re-suspension of sediments following disturbance generally causes rapid release of nutrients to the water column (Jones and Lee 1981; Klump and Martens 1981).

Inorganic nutrient exchanges between the water-sediment interface of intertidal and subtidal sediments have two important, but opposite roles in coastal systems. Release of regenerated nutrients to the water column can supply most of the nitrogen and phosphorus required for phytoplankton primary production (Rizzo 1990; Gómez-Parra & Forja 1993; Cowan et al 1996). In contrast, microbial mats can remove significant amounts of inorganic nutrients from the overlying water column (Teague et al 1988; Ogilvie et al 1997). Removal of nutrients from the water column attenuates the nutrient discharge into coastal waters and contribute to control eutrophication. Since removal and production processes act simultaneously, the direction of net nutrient flux depends on which of these processes dominates.

Dredging is one type of large-scale, anthropogenic disturbance in marine soft-sediment habitats (Lohrer et al 2003). The impacts of dredging have usually been studied with regards to (1) disturbance-recovery of benthic populations and communities, (2) effects of sediment removal and spoil disposal on turbidity and dissolved oxygen concentration (e.g., Brown and Clark 1968; May 1974; Lasalle 1990; Livingston 1996), and (3) the release of toxic compounds following the removal of contaminated sediments (e.g. Windom 1975; Cheung and Wong 1993; Stephens et al 2001). While the potential for biostimulatory nutrient release during dredging has been suggested, there have been very few studies of it (reviewed by Windom 1976; see also Witzig and Day 1983; Michael and Romano 1995).

Lohrer et al (2003) studied the impact of navigation dredging within Debordieu Creek (DC), a temperate estuarine creek in Georgetown County, South Carolina (USA). After collection, a full suite of analyses was performed as part of the NI–WB NERR water chemistry monitoring program. Standard methodological techniques were used during analysis of all constituents. This paper reports results for four variables often considered in assessments of estuarine water quality: total suspended solids >0.7 lm (TSS), orthophosphate (PO_4), ammonium (NH4+), and nitrate-nitrite combined (NOx).

The results show that when water samples were taken synchronously inside the initial mixing zone and beyond the final zone (~720 m downstream), there were measurable, statistically significant differences in TSS, OP, NH4, and NN concentrations, with higher levels in the initial mixing zone in all cases. The biological significance of these effects was probably minimal. Given the natural variability in water chemistry that occurs between years, with the ebb and flood of tides, and with seasonal cycles, the nutrient concentrations measured on 4-Oct-2001 were always well within the range of normal. The ‘‘elevated’’ NH4+ and combined Nitrate-Nitrite (NOx) levels observed inside the dredging plume on that date were at least 50% lower than natural concentrations during a typical annual peak. Furthermore, the levels were approximately 1 order of magnitude lower than the maximum values recorded in DC prior to dredging during NI–WB NERR monitoring from 1998–2001.

To summarize, the effects of the dredging activities on water chemistry in DC appeared to be localized, temporally short-lived, and well-within the range of natural variability in the system at short and long time scales.

The effects of dredging on resuspended sediments and water quality were investigated for the Southern California Contaminated Sediments Task Force (Stivers et al 2004). Their results show that as the numbers and concentration of chemicals in the sediments increases, the potential for effects to aquatic organisms also increases.

Rhoads et al (1995) used a REMOTS registered sediment profile imaging camera for rapid seafloor assessment to define baseline conditions and impacts on sediment quality during both wet (typhoon) and dry seasons in the territorial waters of Hong Kong. Analysis of sediment profile images has allowed assessment of dredging and disposal effects and organic enrichment "hot spots". It appears that tropical storms may have a large effect on mobilization of soft sediments during the wet season, while anthropogenic factors (dredging, disposal, and seabed fishing activities) may assume greater relative importance in some areas during the dry season when severe

storms are less common.

3.2 Effects on Dissolved Oxygen

Physical and chemical alterations associated with dredging include increased levels of suspended sediments and the potential for associated dissolved oxygen reduction and release of natural and industrially-derived chemicals (Lasalle 1990). The magnitude and spatial extent of the suspended sediment field around any dredging operation is a function of the type of dredge used, the physical and biological characteristics of the material being dredged (e.g., density, grain size, organic content) and site-specific hydrological conditions (e.g., currents, water body size/configuration). A generalized "worst case" field can be defined as having suspended sediment levels less than or equal to 500 mg/L at distances less than or equal to 500 m from the dredge, with maximum concentrations generally restricted to the lower water column within 50-100 m, decreasing rapidly with distance. Reduction in dissolved oxygen and chemical release from suspended sediments should be minimal and short lived (Lasalle 1990). The effects of dredging on turbidity are addressed in section 3.4.

High levels of oxygen-depleting substances and nutrients are found wherever very fine-grained sediment deposits and where a great portion originate form waste-water. Mueller et al (1998) discuss the impact of several dredging techniques, and concluded that in order to avoid or mitigate the negative impacts of sediment dredging and relocation on the microbial processes in the water column, cost-efficient measurements and simple forecast calculations may be performed in the early planning phase of such projects. Often the boundary conditions for operations can be fitted in such a way that detrimental ecological impacts are kept within reasonable limits.

Dredging can also alter sediment loading in shallow coastal areas. The effect of increased load of fine sediment on the microbenthos (benthic microalgae, bacteria, and meiofauna) was studied by Wulff et al (1997) in two experiments using undisturbed cores of sandy sediment from a microtidal bay on the Swedish west coast. Within a week, after being covered by fine sediment, benthic microalgae (particularly diatoms) had migrated upward and the oxygen profiles were restored at the sediment surface by photosynthesis. However, the oxygen-producing layer became thinner and the algal composition changed. Bacterial biomass was restored to the same level as in the sandy sediment. Meiofauna also appeared to move upward and the meiofaunal composition was re-established. The results suggest that the microbenthic community of sandy sediment has an inherent capacity to recover after a moderate deposition of fine-particle sediment. Active upward migration of benthic diatoms appears to be a key mechanism for restoring the oxygenation of the sediment surface. The altered sediment type also implies changed species composition, and hence altered benthic trophic interactions, which may affect, for example, flatfish recruitment.

The chemical and biological impacts of anthropogenic physical modifications (including dredging) to tributaries were assessed on New York's Long Island South Shore Estuary by Zakowski et al (2008). Water-quality data collected on Carmans, Patchogue, and Swan Rivers from 1997 to 2005 indicate no significant differences in nutrient levels, temperature, or pH among the rivers, but significant differences in light transmittance, dissolved oxygen (DO), salinity, and sediments were observed. The results of these studies show that Patchogue River (PR) and Swan River (SR) were significantly more saline than Carmans River (CR), PR and SR had less light transmittance than CR, and both exhibited severe warm season hypoxia. CR was rarely hypoxic and only at the lower layer of the deepest station in warm seasons. Deep stations on PR had hypoxic readings year round, but the shallower SR was well-oxygenated at all stations after the fall turnover. There were wide diel and seasonal variations in chlorophyll a on each river, and measurements were significantly higher at poorly flushed stations. In warm seasons, this often resulted in hyperventilation with supersaturated DO in the upper water column on sunny days, and suboxic conditions at nights and/or in deeper layers. PR sediments were anoxic, SR sediments ranged from normal to anoxic, and CR sediments were normal at all stations.

The potential for placed material to leave the disposal site and the impact on water quality of placing 18 million cubic yards (cu yds) of dredge material at one disposal site in Upper Chesapeake Bay (USA) over a five year period were determined through numerical modeling by Johnson et al (2000). The placement method modelled is bottom disposal from a spilt-hull barge. The material is fine-grained maintenance material to be dredged from several upper bay channels. The results of the model runs show that no significant changes were computed in the levels of water column dissolved oxygen at the disposal site after placement of the dredged material.

3.3 Effects on resuspension of bacteria

There is limited information on the resuspension of fecal bacteria in contaminated sediment caused by dredging. The potential exists that a dredging turbidity plume could carry fecal bacteria into shellfish beds. This could result in closure of the beds due to human health hazards. In a similar way, resuspension of sediment-bound pathogenic bacteria can result on an impact on the health of swimmers or other users (http://www.nj.gov/dep/cmp/final_analysis.pdf ).

Grimes (1975, 1980) studied fecal coliform (FC) concentrations in the immediate vicinity of a maintenance dredging operation in the Mississippi River navigation channel. These increased significantly (F test) and increased counts were attributed to the disturbance and relocation of bottom sediments by dredging and a concomitant release of sediment-bound fecal coliforms.

Presumably, if dredging promotes resuspension of sediment bound FC, it should resuspend all sedimented bacteria, including other fecal indicators and enteric pathogens. However, this supposition had not been documented in published literature. For this reason, a study was undertaken in the summer of 1976 to determine the bacteriological effects of hydraulically dredging bottom sediment known to be heavily contaminated with metropolitan sewage effluent. The results of this study suggest that neither turbidity nor bacteriological effects of dredging extended far downstream. Within less than 2 km below the dredge spoil discharge area, the river had recovered from the effects of dredging. In fact, data suggest that water quality 2 km downstream and beyond became progressively better than upstream water quality. This was probably due to natural sedimentation of suspended materials. However, it is possible that dredge-suspended particles could have served as new

adsorptive surfaces in the water column, thereby increasing the rate of adsorption or flocculation (with subsequent sedimentation) of normal suspended, planktonic (unattached or epipsommic) indicator bacteria.

Babinchak et al (1977) studied the effects of offshore dredge spoil deposition on bottom water quality and on bottom sediments at the offshore disposal site. Material (about 1.2 million m3) was removed from the Thames River (USA) by bucket dredging and transported, via hopper barges, to an offshore disposal site in Long Island Sound. MPN FC indices in the top 1 cm of sediment from the dredging site averaged 14,000 (n = 5) FC per 100 ml of sediment before dredging. Deposition of this material had no significant effect on MPN FC indices in either bottom water samples or bottom sediment samples collected from stations in the spoil deposition area. They attributed this lack of effect to the dilution of surface, bacteria-laden sediment with much greater quantities of subsurface, bacteria-free sediment. Unfortunately, sediment profiles (cores) were not analyzed to substantiate this hypothesis.

3.4 Effects of Turbidity

The information available on aquatic species responses and/or mortality due to dredge induced water quality changes is incomplete. It is known however that egg and larval forms of aquatic biota are more sensitive than adult stages. American oyster eggs and larvae are known to be sensitive to turbidity levels and durations that typically occur at mechanical dredging sites. Turbidity is known to block upstream migration of striped bass. Turbidity may therefore, block other anadromous species during upstream migration. Aquatic finfish and blue crabs which winter in the Estuary are lethargic at cold water temperatures. Large-scale mechanical and hydraulic dredging operations could entrain and kill significant numbers, since they would not be able to evacuate a dredging area.

3.4 Summary

While the potential for biostimulatory nutrient release during dredging has been suggested, there have been very few studies of this and associated processes. Dredging activities in estuarine and coastal regions result in the relocation of large volumes of anaerobic porewater and particles, potentially leading to release of nutrients and stimulation of heterotrophic microplankton. Investigations of the extent and the geochemical mechanisms of nutrient release have mainly focused on organic matter mineralization and mineral dissolution during periodic resuspension of surface sediments. While sediment particles may also bind some nutrients, re-suspension of sediments following disturbance also causes rapid release of nutrients to the water column. However, some studies suggest that the effects of the dredging activities on water chemistry in coastal waters appear to be localized, temporally short-lived, and well-within the range of natural variability in the system at short and long time scales. Results of modelling studies also suggest that no significant changes in the levels of water column dissolved oxygen take place at disposal sites after placement of significant volumes of dredged material.

There is limited information on the resuspension of fecal bacteria in contaminated sediment caused by dredging, but some studies suggest that neither turbidity nor bacteriological effects of dredging extended more than 2km from the dredging area.

4. Radionuclides in dredged material

4.1 Introduction

A review was conducted to identify publications in both the peer-reviewed and ‘grey’ literature on studies concerned with the assessment of dredging activities involving the potential impact of radionuclides contained within the dredged material. The review was carried out using both electronic search engines (Scopus, Google) and more conventional methods. There was found to be a paucity of information using these methods, implying that such specific assessments are not frequent or widespread. The best documented relevant studies were international initiatives, carried out at the behest of the London Convention, those carried out by or on behalf of national regulatory bodies and those carried out at a European level.

4.2 Regulatory background

All estuarine and marine sediments are radioactive to different degrees. The contributing radionuclides may be: entirely of natural origin, in ‘background’ concentrations; naturally-occurring radionuclides in enhanced concentrations due to human activities (TNORM – Technologically-enhanced Naturally Occurring Radioactive Material); and, radionuclides introduced to the environment as a result of nuclear power generation or nuclear weapons production and use. In some circumstances radionuclides from several categories may co-exist. Estuarine and coastal sediments may contain elevated levels of radionuclides as a result of human activity, in common with other contaminants, and can sometimes reveal a history of local contamination (e.g. Kershaw et al 1990).

The London Convention 1972 was modified in the early 1990s to ban the sea disposal of all radioactive wastes. This superseded an earlier ban that applied to high-level radioactive waste only. It immediately became apparent that some further consideration was required to take account of the ubiquitous nature of radioactivity, from both natural and artificial sources. For example, a strict interpretation of the Convention would have prohibited the disposal of any materials (including the practice of burial at sea of human cadavers). As a result the International Atomic Energy Agency was requested to develop the concept of de minimis, to define concentrations of radionuclides that were sufficiently low to allow their disposal within the LC.

The IAEA published a consultation document introducing the de minimis concept in 1999 (IAEA, 1999), and this was further revised to provide an agreed radiological assessment procedure (IAEA, 2003). This was based on the ICRP (International Commission on Radiological Protection) principle, accepted by the international community at that time, that protection of the human population would ensure adequate protection of non-human populations. More recently this philosophy has been challenged and efforts have been made to devise an agreed methodology to estimate doses to non-humans. However, at the present time there is no process that has been agreed at an international level, although efforts have been made in a number of countries (including the UK by the Environment Agency) to formulate an appropriate assessment framework. At a European level a series of programmes have been funded by Euratom to develop assessment databases and procedures, including the current PROTECT programme (Protection of the environment from ionising radiation in a regulatory context: http://www.ceh.ac.uk/PROTECT/pages/env_protect_radio.html). Within England and Wales the Environment Agency has published 2 reports describing recommended assessment strategies that will be required under the Habitats Regulations (Copplestone et al 2001; Copplestone et al 2003). This will be relevant to any dredging operations that have an impact on protected sites designated under these regulations. However, under the LC the concept of de minimis is based on the protection of human health.

4.3 Potential impact of radionuclides associated with dredging activities

Under the LC any dredged material containing radioactivity at de minimis levels can be disposed of. Under Part II of the Food and Environment Protection Act 1985 (FEPA), the appropriate licensing authorities in the UK are required to licence that activity, requiring an assessment to made in situations in which compliance with de minimis levels has not been established or may be exceeded. A review was conducted in 2006 of the implementation of de minimis for dredging operations in England and Wales (Cefas 2006). In most regions radionuclide concentrations fell well within de minimis criteria. However, dredging operations in the Irish Sea included material that approached the de minimis criteria, particularly in terms of estimated radiation dose to the crew of the dredger. The report made a number of recommendations intended to increase the reliability of the dose estimates, including direct dose measurements and more realistic exposure data. These higher doses at ports such as Mayport on the Cumbrian coast, were due to a combination of artificial radionuclides released from the Sellafield nuclear fuel reprocessing plant (especially 137Cs, 241Am, Pu isotopes) and enhanced levels of naturally-occurring radionuclides resulting from non-nuclear industry (phosphate processing) near Whitehaven (Poole et al 1995).

5. The impacts and monitoring of maintenance dredging - related to turbidity

5.1 Introduction

The Marine and Coastal, Access Act 2009 provides statutory control of the disposal of waste material to sea from ships in UK waters. This disposal is predominantly dredged material, a large proportion of which arises from the maintenance dredging of port and harbour facilities and their approach channels (Bolam et al 2006).

Turbidity is an index of the scattering of light by suspended particles as it passes through water (Davies-Colley and Smith 2001). The principal cause of dredging-related turbidity is the presence of sediment suspended in the water column. The effect of suspended sediment on turbidity is dependent on the sediment particle size and the sediment concentration. Increasing sediment concentration (without affecting sediment size) increases turbidity. Fine sediment such as clays and silts generate much higher turbidity than a similar concentration of coarse sediment. The environmental effects of dredge-related plumes may result from elevated turbidity (e.g. reduced light penetration) or elevated suspended sediment concentration (e.g. physical abrasion), or a combination of both factors. The nature of the sediment and the hydrodynamic conditions will determine the area over which material in a turbidity plume may be distributed (Bolam and Rees 2003; Erftemeijer and Lewis 2006).

The dredging process typically involves four distinct phases:

1. Dislodging of the in-situ material

2. Raising the material

3. Horizontal transport of the material

4. Placement of the material (Bray 2008).

Elevated turbidity can result from the direct disturbance of the seabed during the dislodging phase. The quantity of material re-suspended depends on the energy applied to the excavation. Fine-grained material will not settle rapidly, causing increased turbidity near the dredging site for and extended period. Turbidity may also become elevated due to the subsequent deliberate or accidental release of sediment into the water column during the raising, transportation or placement phases of operations. During raising and transportation, overflow is the commonest source of turbidity. Placement operation can cause the re-suspension of sediment as dredged material is returned to the seabed (Bray 2008; PIANC 2009).

5.2 Dredging Methods

The type of dredging method and type of equipment employed can have a bearing on turbidity. Dredging methods are discussed in Section 1.2. and are below in outlining turbidity impacts.

5.2.1 Hydraulic Dredging

For maintenance dredging, where sediments tend to be relatively unconsolidated, TSHDs are most commonly used. CSDs are more usually employed in areas with more consolidated bottom sediments (PIANC 2009, Bray 2008).

TSHDs are self propelled vessels with a hopper for sediment storage. The hopper typically has a capacity of 750 to 10,000 m3. One or two suction pipes connect the vessel with the draghead(s). The draghead(s) are lowered to the seabed where they move along the seabed (PIANC 2009; Bray et al 1996).

A slurry of water and sediment is pumped up the pipes into the vessel’s hopper. TSHDs unload sediment through doors or valves in the bottom of the hopper. Smaller TSHDs may be split-hulled, and most modern TSHDs have the ability to discharge their sediment cargoes using pumps and pipelines, allowing discharge ashore where required. When the slurry of water and sediment arrives in the hopper the dredged material settles and the water drains off through a hopper overflow system. As the rate of settlement is principally dependent upon grain size, the overflowing water may contain a significant amount of sediment, especially of finer grain sizes. The presence of this sediment can result in a significant turbidity plume (PIANC 2009).

Other sources of turbidity associated with TSHDs are the hopper overflow, turbulence caused by the vessel’s propeller, and sediment escaping through the vessel’s intake bypass. Intake bypasses are fitted to TSHDs to prevent water being discharged into the hopper at the commencement and conclusion of dredging. These systems may be known as “lean mixture overboard” (LMOB) systems (John et al 2000) or “automatic light mixture overboard” (ALMOB) systems (Bray at al 1996). Overflow of the hopper may occur accidentally or deliberately. Economic benefits can result from deliberate use of the overflow when dredging in sandy conditions. Rapid settlement of relatively coarse sands in the hopper leads to a large sediment concentration difference between the intake and overflow, leading to a greater final sediment load in the hopper. When dredging on a seabed dominated by fine sediments, slow sediment settling rates in the hopper lead to negligible differences between the sediment concentrations in the intake and overflow, removing any benefit of the overflow’s use. Use of the overflow in fine sediment conditions is likely to lead to elevated turbidity and visible turbidity plumes may result (Bray et al 1996; Bray 2008; PIANC 2009).

High turbidity may also result if pump - or sidecasting-discharge methods are used to empty the hopper. Both of these methods involve the pressurised pumping of the hopper slurry. Pump-discharge involves the ejection of the slurry from a nozzle mounted on the vessel. Material can be spayed up to 100m from vessel through the air. This material may be used for beach discharge, but if the jet is landing in water then high turbidity is likely to result. Sidecast-discharge is a method used in specially designed dredger that do not have hoppers. Instead, discharge occurs through a side-mounted boom (up to 90m long) directly back into the water. Dredgers of this type are used for the maintenance dredging of long navigational channels where off-site disposal would be uneconomic. High levels of turbidity can be induced (Bray et al 1996).

CSDs are used mainly for capital dredging projects (Bray 2008). They typically consist of a pontoon equipped with a rotating cutter head and an adjacent suction pipe that collects a mixture of water and cuttings. They differ from TSHDs in that dredging takes place while the vessel is stationary. The mixture of water and cuttings is pumped through a discharge pipeline to its placement location or a storage facility such as a barge. CSDs are best suited to the removal of hard sediments and are therefore rarely used for maintenance dredging. The economics of CSD operations depend on efficiency. High dredging efficiency is closely linked to low turbidity at the cutter head. It is therefore uncommon for turbidity generated by the cutter head to cause environmental concern (John et al 2000). Sediment spillage and resultant turbidity elevation during loading operations can be reduced by avoiding the overloading of barges and hoppers (PIANC 2009).

5.2.2 Mechanical dredgers

Mechanical dredgers of many different types exist, normally discharging into independent hopper barges.

Grab dredgers operate using a crane-mounted grab which is lowered to the seabed to capture sediment. This method may cause minimal disturbance and dilution of clays compared to the hydraulic methods employed by TSHDs and CSDs, but may cause high turbidity in loose silts. A significant fraction of the sediment being lifted may wash out of the grab as it is hauled through the water (PIANC 2009). Closed grabs (though rarely used) can reduce the generation of suspended sediments (Bray 2008).

Backhoe dredgers employ an hydraulic excavator in place of a crane-mounted grab, but otherwise operate similarly to grab dredgers. Like grab dredgers, sediment can wash out of the bucket as it moves through the water, resulting in elevated turbidity. However, they can be fitted with closed buckets to minimise sediment loss (Bray 2008).

5.2.3 Hydrodynamic Dredging

Hydrodynamic dredging may be achieved using the methods of agitation dredging, water injection dredging or underwater plough or sweep dredging.

Agitation dredging is most applicable to the maintenance dredging of fine sediments. The seabed sediments are disturbed using water jets, raking or pumping. Once suspended, these sediments are then re-located by naturally occurring hydrodynamic processes. It is clearly necessary that sufficient understanding of the hydrodynamics exists in order that the fate of re-suspended sediment can be predicted (Bray et al 1996). It is clear that considerable turbidity may be generated using this dredging method – indeed the method relies specifically on the generation of a turbidity plume to re-locate sediment. Agitation dredging is most suited to silt and fine sand sediments such as those which typically accumulate in navigation channels (Herbich 2000). The high levels of turbidity produced mean that agitation may be less suitable for environmentally sensitive projects (Bray 2008).

A variation of agitation dredging is water injection dredging. A fixed array of nozzles is lowered from a vessel until they penetrate the seabed. Low-pressure water is injected into the near-surface sediments until the density of the sediment is reduced to the point at which gravitational forces induce sediment flow (Bray et al 1996). During water injection dredging sediment is not usually re-suspended, and so the technique tends to generate relatively little turbidity (PIANC 2009). The technique is mainly used for dredging in tidal basins where significant natural accumulation occurs. As most of the re-suspended sediment is close to the bed, the effects on turbidity in the upper water column is generally limited (Bray 2008).

Underwater plough or sweepbeam dredging employs a large steel bar or box suspended horizontally from a vessel, or a barge towed by a tug. The bar or box is dragged across the seabed to level areas that have been previously dredged, or to move sediment from restricted to more accessible areas for removal by a TSHD (PIANC 2009; Bray et al 1996). A cloud of suspended material is produced ahead of the plough or sweepbeam, but most of this material remains close to the bed (Bray 2008).

During the excavation phase of dredging, SDs and CSDs generally produce the least turbidity effects, followed by TSHDs, mechanical dredgers and TSHDs when operated with overflow. In the placement phase mechanical dredgers produce less turbidity than CSDs and TSHDs as the dredged material is less disturbed. It therefore follows that hydraulic dredgers are to be preferred if the dredging area is sensitive to turbidity impacts. If the placement are is sensitive, mechanical dredgers may be the preferred option (Bray 2008; PIANC 2009).

5.3 Turbidity Impacts

Dredging releases sediment into the water column, forming a sediment plume, of which there are two types, or phases; the dynamic phase (where the plume moves under its own mass or momentum) and the passive phase (where the plume moves due external influences acting upon it). In the dynamic phase, plume behaviour is mainly determined by the nature and concentration of the material and how it is placed into the water. In the passive phase, plume movement is controlled to a greater extent by the strength and direction of the current (John et al 2000).

The sources of sediment plumes are essentially the losses, deliberate and otherwise, that occur during a dredging operation. There are three primary influences on the generation of sediment plumes: the dredging operation, the material, and the hydrodynamic conditions in which the dredging takes place (John et al 2000).

Fluidised clays and other fine sediments from CSDs and TSHDs can cover excessive areas when unconfined. Fluidised sediments may take a considerable time to consolidate, providing an ongoing source of turbidity following placement (PIANC 2009).

When considering the impacts of dredging-related turbidity it is important to consider the sensitivity of the environment. Where high levels of naturally-occurring turbidity occur (possibly as a result of rapid tidal flows or storm-related high wave energy events) the turbidity and suspended sediment concentration may be orders of magnitude greater than those effects generated by dredging activity. Fauna in habitats with high levels of turbidity disturbance are likely to be adapted to cope with such events, and the habitat is therefore likely to be less sensitive to negative effects. Typical benthic invertebrates in these environments are small, surface-dwelling, fast-growing species with the ability to re-colonise disturbed areas rapidly. Areas where high levels of turbidity are rare, such as seagrass beds or coral reefs, are likely to be far more sensitive to such disturbances. (PIANC 2009; Herbich 2000; Erftemeijer and Lewis 2006).

The physical impacts of dredging operations and their effects on the environment are considered in PIANC (2009). Table 5-1 in this report summarises physical changes and their environmental effects. A section of this table dealing specifically with the re-suspension of sediment is reproduced below:

Table 1 – The physical impacts of sediment re-suspension and their environmental impacts (from Table 5-1, PIANC (2009)).

Physical change

Potential environmental effect

Examples of impact

Re-suspension of sediment matrix into water column

Release of particulate matter

Behavioural / physiological responses to increased suspended solids (e.g. physical abrasion, visual effect of plume, effect on water intakes)

Reduced light penetration

Behavioural / physiological responses to increased turbidity (e.g. loss of growth for eelgrass beds, reduction in primary production for phytoplankton)

Release of nutrients

Behavioural / physiological responses to enrichment (e.g. algal blooms)

Release of toxic chemicals

Behavioural / physiological responses to contaminants (e.g. bioaccumulation of metals in fish)

Release of organic matter

Behavioural / physiological responses to oxygen depletion

User conflicts

E.g. aesthetics, diving, fishing

Traditional fears of water quality degradation resulting from the re-suspension of sediment during dredging and placement operations are mostly unfounded (Herbich 2000).

5.4 Methods of Reducing Dredge-related Turbidity

There are numerous management practices that can be employed to reduce the impacts of dredging operations. These management practices are detailed and discussed in PIANC (2009).

There are several technical practices that can be employed to reduce turbidity impacts associated with dredging operation.

Careful management of TSHD overflow systems should be employed to minimise turbidity. Most modern TSHDs discharge overflow at the level of the vessel’s keel. Some TSHDs can be fitted with a ‘Green valve” device in the overflow system designed to reduce turbidity. The device avoids the entrainment of air in the overflow mixture so that once released below the keel the mixture descends rapidly to the seabed (PIANC 2009; Opuji and Ishimatsu 1975). Turbidity associated with the vessel’s propeller may be reduced by careful navigation. Excess water can be released close to the draghead, or even re-used in the drag-head jets (PIANC 2009).

Some dredgers have been modified specifically to produce low levels of turbidity during dredging and placement. Examples include: environmental disc bottom cutter dredgers, sweep/scoop dredgers, environmental auger dredgers, cable arm clamshell dredgers and horizontal profiling grab dredgers (PIANC, 2009).

During transportation spillage through from hoppers during transport should be negligible (or zero) provided bottom doors and seals are well maintained. Well maintained hydraulic pipelines will have negligible or zero losses. Placement of fine material (a feature of navigational maintenance dredging) will inevitably result in a temporary, localised increase in turbidity (PIANC 2009).

Silt screens or curtains may be used to control the initial dispersal of a sediment plume. This technique may be used during dredging or placement. The screens can either hang from surface floats or stand, attached to the seabed and held upright by sub-surface floats, or a combination of the two methods may be employed. The applicability and effectiveness of this type of measure depend on a number of factors. The location and type of dredging (or placement) operation must be considered, as well as the prevailing wave and current conditions. Screens may be used to completely enclose an operational area, or partial screening may be employed to provide protection to a sensitive area. Some water flow will always occur under or through the screens, carrying with it some sediment load. Such screens therefore can never provide an impermeable barrier to turbidity, though they can be highly effective. Significant current speeds or waves can render the screens too difficult to handle, precluding their use. Current speeds in excess of 0.5 ms-1 are likely to render silt screens ineffective. Difficulties may arise during removal of the screens as a result of the mass of sediment attached. The mass of sediment attached to the screens may even prevent their removal, or cause substantial turbidity plumes during removal operations, and the presence of the screens may present a navigational hazard to vessels engaged in the dredging operations (PIANC 2009; Bray et al 1996).

Air bubble screens do not interfere with logistics (unlike silt screens) but have limited effectiveness and may only be applicable in certain situations such as well constrained channels. They have high power consumption (PIANC 2009).

Materials transported hydraulically and directly placed in an open water site can cause significant turbidity plumes when placement is uncontrolled. Turbidity can be reduced by employing low discharge velocities and releasing sediment close to the seabed. Discharge velocities can be reduced by using a diffuser on the end of the discharge pipeline (PIANC 2009).

5.5 Turbidity Monitoring

A need for monitoring turbidity may be identified during the pre-operational environmental assessment (to ascertain background conditions and their variability) or during dredging operations themselves (to determine the dredge-related turbidity levels). Turbidity monitoring can be achieved using a variety of methods, discussed briefly below.

5.5.1 Water sampling

Known volumes of water samples can be filtered and the mass of the residue determined. The SSC can then be determined in terms of a mass of sediment per a unit volume of water. This method provides an accurate measurement, but has the drawback of providing a single data point in time and space. Spatial accuracy of sampling (especially sample depth) may be difficult to achieve, and this may lead to misleading data, especially where strong vertical gradients of SSC exist.

5.5.2 Sediment sampling

Sediment traps can be deployed on the seabed to capture sediments as they settle, or traps can be suspended in the water column and be designed to interrupt the flow and cause suspended sediments to be collected. A common example of the this second type is the Booner tube developed by the University of Sussex (Marion et al 2005). Estimates of the sampling efficiency can then be made in order to estimate the flux of suspended sediment (e.g. Hargrave et al 1979; Marion et al 2005).

5.5.3 Optical monitoring

Optical instruments can measure turbidity by monitoring optical backscatter (OBS) (e.g. Seapoint OBS sensors) or transmission (e.g. Sequoia instruments LISST instruments). OBS instruments are more sensitive to fine sediments (14-170μm) in suspension than acoustic instruments (Creed et al 2001). Optical instruments can record time series of turbidity, revealing elevated turbidity signal that may be related to dredging or disposal operations. Although they can provide this temporal spread, they are typically deployed in a fixed location and so a single instrument cannot provide spatial data. If spatial data are required, more than one optical instrument may need to be deployed. It is important to locate optical instruments in the correct place to ensure that the turbidity plume is actually sampled. Determining the optimum location will require some knowledge of the hydrodynamic conditions in the monitoring area.

Optical instruments that employ more than one optical frequency (e.g. the LISST 100X) can distinguish between changes in sediment particle size and concentration. They therefore don not require in-situ calibrations and can be used to establish calibrations for single frequency optical sensors if co-deployed.

A drawback of many OBS instruments is that only relative turbidity is recorded. Turbidity is a complex property, responding to both the sediment concentration and the sediment particle size distribution. Calibration will be required if OBS response is to be converted into an SSC. Such a calibration will require measurement of in-situ SSC that can be compared with OBS so that the relationship can be characterised. The necessary SSC data may be acquired by analysing water samples, comparing OBS with in-situ SSC measurements (for example using a LISST 100X) or by characterising the OBS response to a series of known SSC concentration under laboratory conditions. It is vital that the sediment type used in the laboratory closely match those found in-situ. If suspended sediments are heterogeneous, calibration can be achieved by assessing the optical properties of the fraction separately using laboratory techniques. The separate fraction results can then be interpolated to provide an in-situ calibration (Bunt et al 1999).

5.5.4 Acoustic monitoring

Acoustic monitoring of turbidity may be achieved using instruments based upon acoustic backscatter. An increased concentration of suspended sediments leads to an increase in the backscattered acoustic energy. Acoustic instruments are more sensitive to coarse (75-250μm) sediments in suspension (Creed et al 2001) and are therefore of relatively little use in monitoring turbidity from maintenance dredging (which tends to consist of fine sediments). Acoustic instruments can be deployed in-situ (e.g. Wall et al 2008), but an interesting application is to mount the instrument looking downwards on a vessel. The vessel can move through an area and measure a turbidity profile through the water column (e.g. Wood and Boye 2007). It should be noted that only relative turbidity can be measured without further work to calibrate the acoustic backscatter intensity against SSC (e.g. Creed et al 2001).

6. Capital dredging impacts

6.1 Introduction

Capital dredging generally involves the removal of ‘virgin’ material that is relatively stable and has become consolidated under the existing hydraulic regime. However it also includes the removal of material from previously dredged areas where sedimentation has since occurred and has not been disturbed by further dredging over a period of time (10 years). In such cases consolidation of the deposited material occurs and the physical properties of the bed will revert to similar characteristics to the virgin material and is therefore treated as capital dredged material.

Unlike maintenance dredging which is removal recently deposited material chemical contamination associated with capital dredged material is most likely to be from historic rather than current anthropogenic inputs as it is undisturbed, consolidated material. The type and extent of any contamination will vary depending on the location due to human activities and geology.

Capital dredging can also involve the uses of explosives in the dredging of hard rock. The explosives are placed in boreholes drilled into the rock, the work is usually carried out from a floating jack-up pontoon. After blasting the rock is removed by bucket, grab or cutter suction dredgers. Considerable resuspension can be associated with hard rock dredging (Dearnaley et al 1996). Noise effects on fish and mammals from the blasting are another concern.

The previous sections outlined impacts associated with maintenance dredging and these impacts can also occur with capital dredging due to the range of material involved. However as capital dredging is usually associated with removal from areas that have previously been undisturbed it can have additional impacts, these are outlined below.

6.2 Physical effects

Changes in channel shape and dimensions resulting from capital dredging operations in an estuary may permit a saltwedge intrusion to travel further upstream than previously, increase shoreline wave action, change tidal range, tidal currents, suspended sediment load, including the creation of a sediment trap (Dearnaley et al 1996), and suspended sedimentation in areas away from the deepened part of the river. The hydrodynamic changes and their effect on sediment erosion, deposition and transport may cause secondary geomorphological changes away from the dredge location, including the potential erosion of intertidal areas (Dearnaley et al 1996). These processes are affected by the seabed sediment characteristics, underlying geology and particularly on mudflats, the flora and fauna (Dredging and disposal: Changes to hydrodynamic regime and geomorphology. www.ukmarinesac.org.uk/activities/ports/ph5_2_7.htm).

The removal of sediment from the transport system could affect erosion and sedimentation processes and ultimately affect the form of the estuary, possibly depriving downstream coastal areas of sediment required to maintain coastal stability. Equally if the sediment is placed back within the same system, although the net change is insignificant the location of the sediment may change promoting additional siltation in specific areas. By contrast, careful design of disposal can result in intertidal areas being increased (Dredging and disposal: Changes to hydrodynamic regime and geomorphology. www.ukmarinesac.org.uk/activities/ports/ph5_2_7.htm).

Under some conditions slumping of the seabed into the dredged areas may occur and may lead to smothering. Such slumping is common with the construction or deepening of navigation channels, where often it is not possible to establish stable channel side slopes during the initial capital dredge (Dearnaley et al 1996). Additionally, there may be changes to sediment type after dredging (Bray (Ed) 2008).

In cases where the capital dredging is being undertaken for purposes of improved navigation and development of port facilities the dredging is, more often than not, likely to lead to an increasing requirement for maintenance dredging. Which will in turn lead to an increase in the long-term release of bed material into the water column (Dearnaley et al 1996).

An example of effects as a result of capital dredging can be seen at Harwich Harbour. Major channel deepening works in the approach to Harwich Harbour has altered the sediment transport regime. The capital dredge increased siltation in the harbour, which subsequently reduced the amount of sediment input into the Stour/Orwell Estuaries and increased the requirement for maintenance dredging. In this case the capital dredge has created the conditions for increased erosion, which is sustained by the regular removal of sediment from the harbour for disposal at sea (Dredging and disposal: Changes to hydrodynamic regime and geomorphology. www.ukmarinesac.org.uk/activities/ports/ph5_2_7.htm).

6.3 Chemical effects

Isolated deep depressions can occur and may result in stagnant and oxygen depleted water. Also in an estuary or bay, the dredged channel can allow saltwater to move further into the estuary than would normally occur, increasing the salinity and shifting the ecosystem to a more marine environment (Herbich 2000; Bray (Ed) 2008).

6.5 Biological effects

One of the most obvious effects of capital dredging are habitat and feeding ground loss through the removal of material (St Lawrence Centre 1993; Dredging and disposal: Changes to hydrodynamic regime and geomorphology. www.ukmarinesac.org.uk/activities/ports/ph5_2_7.htm). This is generally more of an issue for capital projects as it is involves removal of previously undisturbed material. Additionally, organisms within the removed material will be destroyed and smothering by resuspended material will impact the remaining organisms. Also the alteration of hydrodynamic conditions can alter fish movements and can lead to an alteration of species composition i.e. more marine than brackish species.

6.6 Noise effects

Noise effects can be more of an issue with capital dredging projects because of the consolidated nature of the material to be removed and CSD and BLD which may be employed can generate a significant noise. These dredgers can be a continuous source of significant noise levels, reaching 100 to 115dB in the immediate vicinity of the dredger. This noise diminishes to acceptable levels (50-70dB) a few hundred metres from the dredging site (Bray 2008).

Rock blasting undertaken at Holyhead in Wales, was limited to minimise the disturbance to breeding birds in April and July inclusive, otherwise known as an ‘Environmental window’ (Eisma (Ed) 2006).

6.7 Socio-economic effects

There may be conflicts with other users (fisheries and recreation) such as changes to shipping lanes resulting in displacement of vessels. However, there may be benefits of such works, such as expansion of the harbour or port bringing additional revenue and jobs to the area.

6.8 Cumulative/in –combination effects

There may be in-combination affects associated with other projects within the area, such as maintenance dredging, and this must be assessed on an individual basis.

6.9 Gaps in knowledge

Many of the issues associated with capital dredging are shared with maintenance dredging. However gaps in our knowledge include:

· The noise, both above and below the water line, associated with cutter suction dredgers, bucket ladder dredgers and the blasting from rock removal and their impacts on fish and mammals

· The effects of the dredge in the immediate vicinity of the activity are relatively easy to predict and quantify but the far field effects (in particular estuaries) are more difficult.

· The changes relative to background conditions (i.e. ‘natural’ morphological change) are difficult to quantify such as identifying/establishing cause and effect between far field and dredging activity difficult.

· Modelling capabilities for morphological change such as adequate validation data, capability of model to account for seasonal variations, adequate model runs, saltation (end effect point).

· Quantification of cumulative effects.

7. Mathematical models used to assess the impact of dredging activity

7.1 Introduction

The impact of dredging activities is mainly attributable to two mechanisms: the removal of benthic substratum and the suspension and subsequent deposition of material rejected by screening or overflowing from the hopper (Newell et al 1998). The impacts of these