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Mitigation of diffuse agricultural pollution using buffer zones, ditches, ponds and subsurface bioreactors – Review

Jane Hawkins and Martin BlackwellRothamsted Research, North Wyke, Okehampton, Devon, EX20 2SB, UK

IntroductionAgricultural land is widely acknowledged as being a major source of environmental

contaminants such as nutrients (especially nitrate (NO3) and phosphorus (P)), pathogens,

pesticides and sediment that contribute to diffuse pollution that can lead to contamination of

surface water bodies. In order to reach the target ecological status for UK waters as outlined

in the Water Framework Directive (WFD) (2000/60/EC), a major requirement is the reduction

or mitigation of DWPA. In order to assist with this the UK government Department for

Environment, Food and Rural Affairs (DEFRA) has established a Catchment Sensitive

Farming (CSF) initiative, the principle aim of which is to raise awareness of DWPA, and to

encourage voluntary action by farmers to adopt measures to reduce transport of pollutants.

A total of 44 mitigation methods to control DWPA have been identified by Cuttle et al.

(2007). One of the methods is the establishment of riparian buffer zones to intercept the

transfer of pollutants to watercourses. These have be shown to be the most cost effective

approach for the reduction of P transfer from agricultural land to surface waters, compared

with a range of other mitigation methods (Haygarth et al., 2009).

During the past few decades the quantity of published literature on buffer zones and their

functioning and design has steadily increased, but much of the research has failed to adopt

a multi-functional approach, instead focussing on specific issues such as sediment and P in

isolation (Dorioz et al., 2006; Owens et al., 2007). Furthermore, there is still inadequate

understanding of many of the basic mechanisms and processes involved in buffer zone

functioning and in particular the compatibility of the different processes that control the

various forms of DWPA which buffers can mitigate. This has resulted in their inappropriate

and inefficient application and buffer zones have been found not to deliver their anticipated

benefits to UK water quality (Leeds-Harrison et al., 1996). Even with protocols for the

implementation of buffer zones having been described (Environment-Agency, 1996; MAFF,

1997), and although they are included in current environmental management schemes, a

strategic implementation policy with regard to buffer zone establishment for protection of

surface waters has failed to be developed. One of the failings of the protocol described by

these organisations has been their ‘broad-brush’ approach, which, in a country where buffer

strip width is largely constrained by field and farm size compared with, for example, North

America, is often impracticable. No consideration has been given to individual location

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characteristics, which has often resulted in inefficient buffer zones being established. What is

required, is a more strategic, targeted, and integrated approach, which would not only

deliver better water quality benefits, but more efficient use of land and landscape features.

One of the main problems in the UK with regard to buffer zone efficiency is the hydrological

by-passing or short circuiting of buffer zones via agricultural runoff, channelized flow and

sub-surface drainage (Leeds-Harrison et al., 1996).

Little work has been carried out on the potential role of other complementary edge-of-field

mitigation methods such as managed ditches and ponds, which may improve pollutant

removal, especially if used either alongside or instead of the more conventional use of buffer

zones. By their very nature, ditches are well placed to interact with a large proportion of

water moving from agricultural land to rivers and lakes. Simple techniques can be applied

that enable ditches to act as linear wetlands, providing the capability of water quality

improvement/pollution control together with hydrological regulation (Posthumus et al., 2008).

Evidence suggests that farm ponds could also be effective at reducing the nutrient loading of

surface waters draining from agricultural land and can also help reduce the velocity of runoff

waters, thereby helping to reduce downstream sediment losses while at the same time

attenuating flood peaks (Hawkins and Scholefield, 2002; Heathwaite et al., 2005).

Subsurface bioreactors are effectively ditches containing materials with high hydraulic

conductivity and often a high carbon (C) content that are placed below the soil surface in

such a way as to intercept contaminated groundwater, which passes through the reactive

media, and where a range of contaminants including NO3 (Robertson et al., 2005; Schipper

and Vojvodic-Vukovic, 1998) and P (Baker et al., 1998) are transformed into environmentally

benign forms or immobilised. If combined with conventional buffer zones that are subject to

by-passing by subsurface drainage, they could potentially enhance the nutrient removal

capacity of a buffer zone.

This review firstly identifies the main natural processes occurring in edge-of-field mitigation

methods that can be exploited to mitigate the environmental impact of pollutants. Secondly,

the review gives a reflection on the current use and reported efficiencies of buffers zones,

subsurface permeable reactive barriers, ditches and ponds to control DWPA. The main gaps

in our knowledge of how to successfully and optimally implement these measures are

identified.

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Processes mitigating DWPA

Although modern farming practices can often lead to the pollution of surface waters, there

are many generic natural processes occurring within the landscape features such as buffer

zones,ditches and ponds which can be exploited to mitigate the effects of DWPA including.

Table 1. Processes for amelioration of DWPA from Blackwell et. al. 2002

Agriculturally derived pollutant

AmeliorationAssociated water quality problems Natural process for

amelioration

Nutrients (especially N and P) EutrophicationToxicity

DenitrificationPrecipitationPlant uptakeAdsorptionSediment deposition / retention

Pesticides/herbicides Toxicity AdsorptionPlant uptake

SedimentDeposition/Retention

EutrophicationSilting of gravels

Sediment deposition / retentionHydrological regulation

Pathogens DiseaseSediment deposition / retentionAdsorptionPredation

Heavy metals Toxicity

Plant uptakeSediment deposition / retentionPrecipitationSorption

BOD and COD De-oxygenationAdsorptionSediment deposition / retentionOxidation/mineralisation

A brief description and method of mitigation action for these processes is as follows:

A. Denitrification and Nitrification

Denitrification involves the dissimilative reduction of oxidised forms of nitrogen (N), in

particular NO3, to gaseous forms of N, particularly nitrous oxide (N2O) and di-nitrogen (N2)

(Blackmer and Bremner, 1977). Denitrification is carried out mainly by facultative anaerobic

bacteria belonging to a number of genera, which inhabit nearly all known environments

(Groffman, 1994). However, purely chemical mechanisms may result in a similar reduction

of oxidised N-compounds (Brady, 1990; Buresh and Moraghan, 1976). An essential

requirement of the process is anaerobicity, as such conditions stimulate denitrifying

organisms to use NO3 as an electron acceptor in the absence of free oxygen (O2).

Consequently the potential for the process to occur is usually greatest in wet environments.

Most denitrifying bacteria are of a heterotrophic nature and therefore require a supply of

easily oxidisable C as a respiratory substrate (Alexander, 1961). Coupled

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nitrification/denitrification (see oxidation below) can be an important process in some

situations e.g. saturated soils or ponds (Reddy et al., 1989). The particular significance of

this process in terms of nutrient removal is that it involves total export of N from the system

to the atmosphere, rather than temporary storage by other processes such as plant uptake

or organic matter accumulation in the soil.

Denitrification is also an important process for a variety of other reasons; the farming

community is interested in the process as it can lead to large losses of expensive NO3-

fertilisers, with some estimates claiming that up to 30% of applied fertiliser-N can be lost via

the process (Averill and Tiedje, 1981), while environmental managers view the process from

two different perspectives. Firstly, it can perform a useful function by removing NO3- from

diffuse agricultural run-off, and so maintain or improve the quality of surface water bodies

and prevent eutrophication (Blackwell et al., 1999; Woodward et al., 2009; Zaman et al.,

2008). Secondly, it can be viewed as a potentially problematic process, by which the

radiatively active gas N2O is produced, sometimes in large quantities, depending on whether

either partial reduction to N2O or complete reduction to N2 takes place (van den Heuvel et

al., 2009).

In contrast to denitrification, nitrification is the transformation of ammonium (NH4) to NO3.

and dependent on sufficient O2 levels in soil and sediments. Although highly reduced

conditions are favourable for denitrification, anoxic conditions inhibit nitrifiers to an extent

where NO3 becomes limiting. A reduction in denitrification may lead to an accumulation of

NH4 in ponded water bodies. Thus, maintenance of the conditions favourable for the nitrifier

population is important especially where loading of NH4 to waters is already sizeable.

B. Volatilisation

Ammonia volatilisation is a process whereby ammonia (NH3) is produced from ammoniacal

N in solution and is returned to the atmosphere in a gaseous form. The rate of volatilisation

is primarily driven by the free NH3 concentration and temperature (Craggs, 2008). Although

losses of N due to volatilisation are thought to be relatively small (Hargreaves, 1998), levels

ranging from 10-27.4% NH3 have been reported (Banerjee et al., 1990; Gross et al., 2000;

PaezOsuna et al., 1997). This is an undesirable N removal process since NH3 is an

atmospheric pollutant that can have harmful affects on terrestrial and aquatic environments

due to wet and dry deposition (Asman, 1994).

C. Sedimentation

Agricultural land can be a major source of sediment, which in its own right can act as a

pollutant (Owens et al., 2005). Deposition of this material into river bed gravels can cause

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deoxygenation with subsequent loss of invertebrate habitats, and failure of spawning by

salmonids and other fish (Harrod et al., 2002). Nutrients, especially P, as well as pesticides,

heavy metals and pathogens can all be sorbed onto, or associated with, sediment particles,

while high biochemical oxygen demand (BOD) is often related to the presence of organic

particles (Horowitz, 1991; van der Perk and Jetten, 2006). There are two main processes by

which sediment can be removed from agricultural runoff:

1. Flow velocity reduction

This can result from reduction in slope gradient, discharge into ponded water bodies,

infiltration of water into soils or increased friction and detention of surface water by

vegetation, causing surface deposition of particulate matter from suspension (Dillaha et

al., 1989). Densely packed, fine vegetation such as grass offers the most resistance to

flow at shallow depths i.e. has the greatest roughness coefficient (Hook, 2003; Mitsch and

Gosselink, 2000; Munoz-Carpena et al., 1999), though its impact can be reduced in

deeper flows (Hammer, 1992). The greater the reduction in flow velocity or the longer the

period of ponding, the greater the amount of sediment deposited, and in particular finer

particles (Dillaha et al., 1989). The importance of this process for P retention is

demonstrated by Mitsch et al. (1979) who reported P retention by sedimentation of 3.6 g

P m-2 per year in a riparian buffer zone. This was estimated as being eighteen times that

of all other P retention mechanisms.

2. Filtration by vegetation

Vegetation can act as a filter, depending upon its structure and density. Larger soil

particles can become trapped within litter layers or among stems and leaves (Dillaha and

Inamdar, 1997).

D. Chemical precipitation

Phosphorus readily forms precipitates with aluminium (Al), and iron (Fe) under acidic

conditions, while combinations with calcium (Ca) and magnesium (Mg) are more usual under

alkaline conditions (Nriagu, 1972). In low energy environments these precipitates will settle

and become stored. However, anaerobic conditions can have the reverse effect and result

in the mobilisation of previously precipitated P (Patrick and Khalid, 1974), as can alterations

in pH (Richardson, 1989). The relative importance of precipitation processes in preventing P

transfer is not clear because particulate matter of this kind has been shown to be transferred

along various sub-soil pathways (Haygarth et al., 1998), and therefore may simply be

transported in a different form.

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E. Oxidation

In otherwise anaerobic soil environments, some plants can release large amounts of O2 from

their roots, and consequently provide aerobic pockets (Jaynes and Carpenter, 1986). This

can be an important for processes such as the coupling of nitrification and denitrification

processes (Lloyd, 1993; Reddy et al., 1989), optimising the removal of N from the soil.

F. Plant and bacterial uptake

Plant roots remove nutrients from soils by uptake of dissolved solutes in soil water, although

direct exchange between nutrient ions adsorbed onto soil particles and hydrogen (H+) ions

on root surfaces can also take place (Reddy et al., 1989). Where plant growth is strongly

seasonal, the process of nutrient removal and storage by this means is correspondingly

limited to the growing season. During dormant periods, the senescence and decomposition

of above ground plant material results in the release of nutrients which are recycled into the

system. To avoid this, harvesting and removal of the vegetation is required to achieve

nutrient removal (Koopmans et al., 2004). Plant species, age, root architecture and size and

stage of development are important determinants of nutrient uptake (Fohse et al., 1988). In

addition, heavy metals can become incorporated into plant material, as either short- or long-

term storage, depending on plant types and conditions (Klopatek, 1978). Soil bacterial

populations can greatly increase under favourable conditions, resulting in the assimilation of

large quantities of nutrients. However, the storage time can vary considerably, and

unfavourable conditions can result in rapid reductions in bacterial numbers with the release

of large quantities of nutrients (Groffman, 1994).

G. Adsorption

Adsorption of dissolved materials onto soil particles through ionic bonding can be a

significant process for the reduction of concentrations of various nutrients (Leinweber et al.,

2002). Since the capacity for adsorption depends on the amount of soil surface available, the

small size and relatively large surface area of clay particles means that clay soils generally

have the highest adsorption capacity of mineral soils. The type and amount of ions adsorbed

also depends on the anion and cation exchange capacities of a soil which is influenced

largely by the mineralogy of the clay particles (greatest in swelling clays such as smectite),

and the pH of the soil (White and Zelazny, 1986). Cations are adsorbed more strongly than

anions, but at low pH anion adsorption capacity increases. In addition to clays, humified soil

organic material also has a high adsorptive capacity (Stevenson, 1994).

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H. Additional significant processes

Improvements in the quality of water draining from wetlands can result from the inactivation

and predation of pathogenic bacteria and viruses. Protozoa, which are often abundant in

wetlands, can consume large quantities of bacteria, while bacteriophages (viruses that infect

bacteria) can reduce numbers of susceptible bacteria by fatal infection (Kadlec and Knight,

1996; Nuttall et al., 1997).

Different combinations of these processes occur to different degrees in the various features

considered here. It is important to recognise that each type of feature itself may vary

considerably with regard to properties such as slope, soil type, hydrology and vegetation,

and consequently so will the processes acting within them. Here we consider the key types

of mitigation features that can be used to control DWPA.

Buffer zones

‘Buffer zone’ is a generic term referring to naturally or semi-naturally vegetated areas

typically situated between agricultural land and a surface water body, although some buffer

zones can be distal from water bodies (e.g. contour buffer strips – see below). They can be

effective in protecting water bodies and other habitats from harmful impacts such as high

nutrient, pesticide or sediment loadings resulting from land use practices (Blackwell et al.,

1999). Additionally, they act as a physical barrier restricting the spreading of fertilisers and

sprays in the proximity of features along which they are situated. The degree to which

protection is provided by them depends on the processes that can operate in a specific

buffer zone, and this is dependent upon several factors including size (Gergel et al., 2005),

location (Pinay and Burt, 2001), hydrology (Correll, 1997), vegetation (Abu-Zreig et al., 2004)

and soil type (Leeds-Harrison et al., 1996) of the buffer zone, as well as the nature of the

pollution against which it is mitigating. Considerable confusion often arises from the

terminology associated with buffer zones, as many different expressions are employed to

describe a wide range of landscape features used for the mitigation of DWPA. A

comprehensive definition and classification of the various terms associated with buffer

zones, such as buffer strips, contour strips, riparian buffer zones, vegetated filter strips, etc.

is provided by Owens et al. (2007).

The key processes operating in buffer zones varies depending on whether they comprise

freely or poorly draining soils. In buffer zones with poorly draining, wet soils, processes such

as denitrification occur, acting to remove N from the system (Gilliam et al., 1997). The ability

of buffer zones to remove NO3 from agricultural runoff via denitrification has been reported

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by many researchers including Peterjohn and Correll (1984), Cooper (1990), Blackwell et al.

(1999) and Hefting (2006), all of whom report NO3 concentration reductions in surface or

ground water of 75% or more of the original concentration in water discharging into the

buffer zones studied. On the other hand, while P dynamics in buffer zones can be complex,

as described by Dorioz et al. (2006), removal tends to be primarily through the retention of

sediment to which P is bound and consequently is more prevalent in buffer zones with freely

draining soils through which infiltration of runoff occurs, leaving sediment trapped at the soil

surface (Dillaha and Inamdar, 1997; Dillaha et al., 1989). Figures for P removal by buffer

zones are generally impressive with reported retention of total P ranging from 40% to 90%

(Borin et al., 2005; Duchemin and Madjoub, 2004; Schmitt et al., 1999). However, several

researchers report that for dissolved P, buffer zones can sometime be net emitters with

retention ranging from -80% to +95% (Duchemin and Madjoub, 2004; Uusi-Kamppa et al.,

2000). This variability and ability to act as a source is largely attributed to plant uptake of

dissolved P during summer, and its subsequent release due to litterfall in winter. As a

consequence of the different conditions required for optimal performance of the two

processes it is unlikely that N and P removal in buffer zones will occur concurrently to any

great extent in buffer zones. The key processes associated with N removal in buffer zones

usually involve some form of transformation and potential emission (e.g. via gaseous N

emission following denitrification), meaning N removal is effectively sustainable. For P

removal though, concerns exist about the sustainability of the effectiveness of buffer zones,

especially with regard to sediment-associated P, and questions are arising on the effective

life-span of buffer zones for P removal.

Despite the recognised importance of buffer zones in the protection of surface waters and

the publication of manuals recommending their use e.g. Environment Agency (1996), as

such, there has not been any specific policy developed for their implementation in this

context in the UK. The main strategy for implementing buffers in the UK has been through

agri-environment schemes dating back to 1986, generally with an emphasis on biodiversity

and habitat creation. For example, set-aside under the Single Payment Scheme resulted in

the establishment of many buffer zones, although these have not been specifically targeted

at watercourse protection. Set-aside was first introduced in the UK in 1988 (known as the

five-year scheme) (Set Aside Regulations, 1988), as an EU control measure to limit over-

production of cereals and other arable crops, whereby at least 20% of eligible land had to be

taken out of production and managed as rotational or non-rotational fallow. Although set-

aside was introduced as a supply management tool, set-aside has always been recognised

as a potential opportunity for achieving environmental benefits, albeit as a secondary effect

of the production control objective (Evans et al., 1997; Firbank et al., 2003). The percentage

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of land that had to be set aside has been regularly adjusted by the European Commission

according to perceived need, until in the autumn of 2007, the set-aside rate was set at 0%

for the 2008 harvest, in order to increase cereals supply to the market. As a result of this

reduction of set-aside land and potential loss of buffer zones, the UK government have been

reviewing their policy on buffers and how new measures under land management schemes

(Environmental Stewardship) can be implemented to regain the benefits of lost set-aside

land. This has placed buffer zones at the forefront of agricultural policy development in the

UK..

A number of tools have been developed to assist with targeting most efficiently the

establishment of buffer zones. These include the Riparian Ecosystem Management Model

(Lowrance et al., 1998), the Chemicals, Runoff and Erosion from Agricultural management

Systems Model (CREAMS) (Kinsel, 1980) and the Water Erosion Prediction Project (WEPP)

model (Flanagan and Nearing, 1995), all developed in the US. However, these models

require considerable quantities of data, and generally have been developed in landscapes of

different characteristics to that of the UK, where space is of a premium, making these

generally impracticable in the UK. One such tool has been developed by Wood et al. (2007),

optimised for the establishment of buffers for sediment and phosphorus in a UK context. This

tool recommends buffer widths ranging from 2 to 24 metres depending upon soil type and

slope. A summary of efficient buffer zone widths for different functions from studies around

the world is given in Figure 1. This demonstrates the limitations of buffer zones under current

policy in the UK, given that current Environmental Stewardship policy recommends buffers of

widths between only 1 and 6 metres. It should also be recognised that buffer shape and

location can be important, whereby due to hydrological pathways and channelling, key areas

of the landscape may be better targeted as buffers (e.g. Blackwell et al., 1999) rather than a

blanket approach with uniform buffers alongside all watercourses. As well as questions of

effective width, there is still considerable uncertainty about the effective life-span of buffers,

particularly with regard to sediment and P retention. Primary concerns are that buffers could

become potential sources of these pollutants after accruing sediments for long periods, and

interest is growing into how this issue might be addressed through management.

Additionally, as mentioned previously, the multi-functional capacity of buffer zones and their

ability to ameliorate multiple pollutants is unclear, largely as a result of the nature of previous

studies, which have largely focussed on only one or two pollutant types at a time. Interest is

also developing in the ability of buffer zones to remove pathogens from slurry applied to land

(Kay et al.), and other associated pharmaceuticals including Endocrine Disrupting Chemicals

(EDCs) originating from the immunisation of livestock (pers. com. Charles Tyler,

Biosciences, Exeter University). These have the capacity to cause mutations in many forms

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of freshwater fauna (Hutchinson, 2002; Kloas et al., 2009). These topical issues are being

addressed in the project described later in this paper.

Ditches

Ditches are the principal means of land drainage on hydromorphic soils throughout the UK,

and their prime purpose is to remove excess water from agricultural land. In so doing they

constitute one of the main hydrological pathways linking agricultural land and surface water

bodies. Amelioration of diffuse pollution can be effected either at source, through prevention

of mobilisation or via the connectivity/delivery routes to surface water bodies {Haygarth,

2009 #960}. Ideally a pollution control strategy is best targeted at prevention at source, but

realistically, even if best management practices (BMPs) are applied within the first two

stages described above, control within the latter stage will almost always be required. As

discussed earlier control has focused on the establishment of buffer strips, but commonly on

drained land sub-surface drains discharge directly into ditches, resulting in the by-passing of

treatment areas, and the majority of polluted water from agricultural land enters rivers and

streams unaltered. By their very nature ditches are well placed to interact with a large

proportion of water moving from agricultural land to rivers and lakes, but to date ditch

management has focused on maintenance of water transference capacity together with

wildlife and conservation values. Simple techniques can be applied that not only will enable

ditches to provide these benefits, but also effectively to act as linear wetlands, providing the

capability of water quality improvement/pollution control together with hydrological

regulation. The former could help improve water quality through the creation of wetland

habitat supporting processes such as denitrification, sedimentation and plant uptake, while

the latter could help reduce flooding by managing ditches so that they act as multiple

reservoirs, storing water at critical times, and consequently attenuating downstream flood

peaks. To date there has been little published research into the buffering capacity of ditches,

although many anecdotal reports exist. Most studies have focussed on the ecological

benefits of ditches (e.g. Herzon and Helenius, 2008), rather than hydrological regulation and

water quality control. However, the small amount of literature on the potential of ditches to

ameliorate agricultural pollution suggest that in some cases, even without specific

management for these purposes, ditches can remove more than 50% of the inorganic N load

entering them (Kroger et al., 2007) and in excess of 40% inorganic P (Kroger et al., 2008).

This level of performance can be considerably improved as demonstrated by a managed

ditch system coupled with ponds which is reported to have retained ca. 95% of the total N

and P during a single storm event (Chen et al., 2007). Additionally, Lamsodis et al. (2006)

report 30-40% reductions in concentrations of N and P downstream of beaver dams on small

streams and ditches, and it is reasonable to assume that similar or even greater benefits

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could be delivered by artificial dams or baffles, behind which pools and wetlands could

establish. There is some published literature on the water quality benefits of in-ditch

vegetation (Beltman, 1990; Meuleman and Beltman, 1993; Mulholland et al., 2008),

particularly with regard to herbicides and insecticides (Borin et al., 2004; Moore et al., 2001).

Most of these studies have focussed on ditches that have not been specifically managed to

mitigate against DWPA, and the performance of such ditches are highly variable.

Management of ditches for these combined purposes may prove highly effective because

treatment areas would be located in the hydrological conduits themselves. It must be

remembered that ditch management must be tailored to the specific characteristics of

individual sites, which may vary in properties such as slope, width and depth. However, ditch

management techniques applied in the correct fashion could have an important role to play

as part of a holistic approach to catchment management for environmental improvement and

flood prevention.

Ponds

Historically, ponds have been used to treat polluted liquid streams from a variety of settings

including industry (manufacturing process waste) and urban sources (road runoff, municipal

waste and sewage), often as a component of sustainable drainage systems (Woods Ballard,

2007). Various treatment configurations are employed; ponds may be used individually as

settlement tanks, or in a series whereby contaminated effluents are passed through a

sequence of, for example, anaerobic, facultative and maturation ponds which progressively

remove pollutants. They may also be used in conjunction with other systems such as

constructed wetlands as a final ‘polishing’ treatment (Tanner and Sukias, 2003). Due to their

wide application as effective pollutant removers, pond treatment design and technology has

been well documented (Shilton, 2008).

Within agriculture, ponds or lagoons are commonly used to store and treat livestock effluents

such as. slurries and dirty water (Meyer et al., 1997; Nicholson and Brewer, 1997;

Westerman and Bicudo, 2002 ). More recently, the potential of ponds to ameliorate NO3-, P

and sediment from diffuse sources within a catchment setting, is being increasingly

recognised (Fleischer and Joelsson, 2000; Fleischer et al., 1997; Hawkins and Scholefield,

2002). For example, budget studies on a series of restored ponds in southern Sweden has

shown that up to 7000kg N ha –1 yr-1 and 404 kg P ha –1 yr-1 can be removed by ponds

receiving agricultural runoff at the rate of 0.14-5.2 m3 m-2 d-1 (Fleischer et al., 1994). The

efficient retention of nutrients within a pond is dependent on the dynamics of, and the

interaction between the physical, chemical, and biological processes described previously.

Denitrification is reported as the major N removal mechanism from ponds (De Busk et al.,

1983; Fleischer, 1995; Horberg et al., 1991; Yan et al., 1998) and generally takes place in

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anaerobic zones in the sediment at the bottom of the pond. Nitrogen gas diffuses upwards

from the sediment into the pond water column and eventually to the atmosphere.

The majority of N retention is during periods of low flow in the summer and during this period

denitrification may account for up to a 50% loss of the total N removed. Annually,

denitrification may account for 30-40% loss of total N (Jansson et al., 1994) and can be as

large as 90% of the total (external+internal) N load (Fleischer et al., 1997).

There is evidence to suggest that the level of organic N in effluent may increase relative to

that of the influent (Rushton and Bahk, 2001). This is presumably due to transformation of

inorganic N to organic N and subsequent release by the pond biota by exudation or

decaying dead matter, and as a consequence of increasing and maturing pond biomass.

Since organic N is very likely to be in bio-available forms for use by downstream organisms,

it is important to consider organic N as well as inorganic N concentrations when calculating

the N retention and removal efficiencies for ponds. The same may also apply in calculation

of P budgets, where release of organic forms of P should also be taken into account.

The primary mechanism for retention of P in ponds is that of sedimentation and co-

precipitation with calcium carbonate (CaCO3) (Istvanovics et al., 1990) with the size of the

sediment particles a major control on effective retention. Pond soil and sediments strongly

adsorb P and adsorption increases with increasing clay content of the soil and sediment.

The most important removal process of dissolved P in ponds is phytoplankton uptake

(Istvanovics et al., 1990), and uptake generally follows Michaelis-Menten kinetics (Lehman

et al., 1975).

One of the major problems is that during winter time in the UK, nutrient transport and

maximum flow rates are at their greatest, whilst biological loss of N through denitrification is

least effective. This is most apparent in wetland systems and consequently it has been

argued that generally ponds are more efficient than wetlands at nutrient removal, especially

N, since the water retention time and hence residence time of nutrients is generally longer

(Jansson et al., 1994).

The retention time of pond water is the single most critical factor for effective removal of N

and in particular P, since excessive hydrological loading may prevent sedimentation. During

periods of extreme flow, relatively large sand and silt sized particles or aggregates may be

deposited but the smaller clay particles, on which most particulate associated P is attached

may not settle out (Brown et al., 1981). The pond may be overloaded hydrologically to a

point where even silt sized particles are not deposited. However, a system of small ponds

and ditches can help increase efficiency, especially in the case of P removal (Yin and Shan,

2001). Therefore, it is recommended that ponds should be of a sufficient size to allow for

sedimentation even when peak flows occur (Uusi-Kamppa et al., 1997).Recommendations

for the ratio of pond to catchment area range from 1:11 [Yin and Shan 2001] to 1:50 (Uusi-

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Kamppa et al., 1997). Pond depth is an important factor for efficiency and should be

approximately 1m to maintain optimum levels of light and temperature for biological

processes such as photosynthesis and nutrient transformation to take place (Stachowicz et

al., 1994). Furthermore, in the case of N there is an empirical relationship between areal

loading and areal retention of N (kg N ha-1 yr-1) in that an increase in N loading per unit area

can correspond to an increase in areal retention, but a decrease in percentage removal

(Fleischer and Stibe, 1991).

Efficient N removal by denitrification is dependent on the rich supply of organic material

contained in the sediment. Conversely, too much organic material may result in conditions

that are too reduced for optimum rates of denitrification (Fleischer et al., 1997).

Another mechanism for the removal of nutrients is plant uptake by aquatic plants. McBride

and Menzel (1978) published a review on the use of terrestrial or semi-aquatic plants,

aquatic weeds and phytoplankton to recover nutrients from eutrophic farm ponds using

reported kinetics of uptake. Their findings suggest that the higher plants compete well with

phytoplankton for PO43-, but are at a disadvantage for NO3

- uptake. Both NO3- and PO4

3- could

become limiting and that terrestrial plants alone will not remove enough PO43- to prevent

algal blooms. Elsewhere, eight aquatic macrophytes were investigated for their role in

nutrient removal capabilities (Reddy and Busk, 1985). Results showed that removal rates

were generally greater during the summer compared to winter and plant uptake accounted

for 6-75% total N and 12-73% Total P. The results also showed that rates of N uptake

varied between plant species and season. Uptake efficiencies have been reported for a

range of aquatic species e.g. (Horberg et al., 1991; Lavania and Lavania, 2000; McCord and

Loyacano, 1978,De Busk, 1983 #159; Oron et al., 1988; Ouazzani et al., 1995; Polprasert

and Agarwalla, 1994; Zakova et al., 1994).

Ponds offer additional benefits from an ecological perspective, in that in comparison to

ditches, rivers, streams and lakes, ponds can support a high number of species and high

index of rare species (Davies et al., 2009). In addition, ponds can be linked to other forms of

economic output from the farm such as leisure or aquaculture and provide some amenity

value to the landscape (Defra, 2004).

Periodic removal of sediment is required in order to maintain effective biological cycling

since excessive build up of sediment may also eventually lead to the pond becoming too

shallow and may dry out due to evaporation. Furthermore, during the periods when there

may be minimal or no water flow, an inefficient pond could well become a point source of

excessive nutrient concentrations which may lead to potential environmental problems

downstream once flow recommences. Similarly if animal wastes enter a pond during these

periods, there could an excessive and localised build up of pathogenic organisms. In

addition, excessive nutrient loading and the absence of a well structured food web could

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lead to the succession of undesirable or toxic algae blooms (Briand and McCanley, 1978).

However, the methods for disposal of pond sediments which have sequestered large

quantities of nutrients, and in particular P, needs careful consideration since these nutrient

rich sources could be potentially catastrophic to the environment especially if disposed of

next to watercourses.

Other challenges that need to be addressed in pond management include increased N2O

emissions in circumstances where denitrification is incomplete (Martienssen and Schops,

1999). Furthermore, equilibrium reactions of NH4+ with ammonia (NH3

+) leads to some

gaseous loss (volatilisation) of NH3 to the atmosphere particularly in water with a high pH

(Banerjee et al., 1990). Volatilisation also occurs when NH3 inputs exceed the algal NH3

assimilation capacity (Lorenzen et al., 1997). Although losses of N due to volatilisation are

thought to be relatively small (Hargreaves, 1998), levels ranging from 10-27.4% NH3+-N have

been reported (Banerjee et al., 1990; Gross et al., 2000; PaezOsuna et al., 1997).

Regular removal of excessive macrophyte vegetation will be a requirement in order to avoid

infilling of ponds and reduced capacity for storage, but more importantly to reduce the

contribution of decaying material to the water, due to die-back in winter, which can greatly

increase the biological oxygen demand (BOD) and reduce the pond efficiency (Pearson,

2008).

One of the main problems that can arise from nutrient enriched waters is the encouragement

of algal growth, and in particular undesirable toxic cyanobacteria or blue-green algal species.

However, the presence of algae need not be detrimental to a pond system. Indeed they are

recognised as essential to the functioning of water treatment technologies, particularly that of

waste stabilisation ponds (Pearson, 2008). As a result of their photosynthetic activity, algae

provide a valuable function in maintaining the oxygen status of the water thereby promoting

the oxidative decomposition of organic matter by bacteria (Pearson, 2008). Carbon and

other nutrients are produced as a result of decomposition which in turn is utilised by the

algae (Shilton and Walmsley, 2008). It has been suggested that the use of algae can reduce

the required surface area for a pond by as much as a factor of 5 (Picot et al., 1993), however

it is important to note that, as with macrophytes, algal biomass can lead to an increase in

BOD when it dies back and decomposes.

Another consideration is that ponds attract waterfowl, such as migratory geese, that could

lead to additional input of nutrients to waters together with feeding damage to the

surrounding crops (Hill and Frederick, 1997; Post et al., 1998).

Subsurface bioreactors

Subsurface bioreactors, also known as permeable reactive barriers (PRBs), are where

reactive substances such as zinc, zeolites, reactive clays, and chemical oxidants, or high

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carbon containing materials such as activated charcoal, sawdust and compost are placed

below the soil surface in such a way as to intercept contaminated subsurface flows of water.

The materials have a high hydraulic conductivity, providing a flow path through the barrier,

and either transform contaminants such as NO3 by microbial processes into environmentally

benign forms or ‘bind’ them though physical process e.g. P. Choice of material is dependent

on knowledge of the interactive processes between the target pollutant and the material in

the PRB (Yong et al., 2007). The materials should also be resistant to rapid microbial

decomposition so as to avoid frequent replenishment. The materials would need to be

evenly packed into the trench in order to avoid channelling, which would provide preferential

pathways through the barrier and reduced contact time with the material. However, it is

reasonable to expect some degree of channelling due to zones of decomposition over time.

The technology emerged in the 1990’s as an effective way to treat polluted groundwater and,

although applicable for use with any contaminated liquid, they are commonly used to treat

ground water in aquifers (Carey et al., 2002). For example, bark based barrier walls have

been shown to be effective in trapping polycyclic aromatic hydrocarbons (PAHs) in

groundwater with an efficiency range of 77 – 100% (Seo et al., 2007). Application of the

technology for the treatment for agriculturally derived polluted waters is gaining recognition,

particularly in the control of NO3, since high carbon containing materials can provide an

active zone for denitrification. Barriers consisting of a mixture of sawdust and soil in a trench

(35 m long, 1.5 m deep, and 1.5 m wide) reduced groundwater NO3 concentrations of 5-16

mg N l-1 to <2 mg N l-1 (Schipper and Vojvodic-Vukovic, 1998). Similarly, studies using

different sized particles of wood found that up 60% reduction in NO3 concentration levels of

between 23.7 to 35.1 mg l-1 were achievable (van Driel et al., 2006). Recently, Kalin and

Assal (2009) have investigated the use of permeable reactive barriers at catchment scale to

remove nitrate, and results have shown >95% removal efficiency for concentrations up to

100 mg l-1 (Kalin and Assal, 2009). Further studies suggest that that these type of barriers

have the potential to remain effective for up to 10 years, and possibly more, without

replenishment of the reactive material (Blowes et al., 1994; Robertson et al., 2008) An

important consideration is the disposal of potentially nutrient rich material if is removed from

the bioreactor. It may be necessary to process the material e.g. compost, in order to reduce

its nutrient loading.

Each of the features described above are capable of mitigating DWPA in their own right.

However, sometimes the characteristics of a pollution event (e.g. very high rainfall or

pollutant loading), or the nature of the pollutant (e.g. particulate or dissolved) mean that

solitary mitigation features may not be able to manage the pollutants adequately. Therefore

it may prove more sensible to combine features in sequence, thereby creating a multi-

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featured mitigation system which not only provides opportunities for managing a wide range

of DWPA, but also provides a ‘safety-net’ should one or more of the features be ‘breached’.

To date there has been little research into how different combinations of these features may

be used in a farm situation to protect simultaneously against a broad range of DWPA

pollutants, and how effective they might actually be.

Conclusions

This paper has provided an overview of the key issues surrounding the use of buffers for the

control of DWPA. Whilst they are effective in reducing NO3 and can provide an extremely

cost effective option for the control of sediments and P, this is highly dependent on them

being implemented correctly and is site specific.

In addition, there is a lack of robust scientific evidence on their long-term sustainability and

performance under a range of different nutrient loading and climatological scenarios. To this

end, we have provided a brief overview of the nutrient removal capability of other landscape

features such as ponds, ditches and subsurface bioreactors that could be either used in

isolation where buffers are unsuitable, or as part of a suite of mitigation methods, including

buffers, that if used in strategic combinations could enhance pollutant removal and provide a

‘safety-net’ should one component fail. In addition, simple and ‘farmer-friendly’ techniques

applied to ponds and ditches could not only enhance removal of nutrients but also provide a

degree of hydrological control. Although there are issues concerning disposal of nutrient rich

materials from ponds, ditches and subsurface bioreactors, these structures could be viewed

as a way of actively ‘harvesting’ nutrients from polluted waters, and if handled properly, the

materials could provide a valuable source of fertiliser back on the farm.

This review has identified that the areas lacking in the understanding of the efficiency of

buffers and other edge-of field mitigation methods include:

Reduction in buffer efficiency as a result of bypass flow

Temporal and spatial variability in efficiency

Identification of the key indicators of failure in efficiency

Evaluation of mitigation methods as potential point sources

Pollution swapping e.g. N2O, NH3

The management required to maintain effectiveness and sustainability of the

methods e.g. sediment clearance and disposal from ditches and ponds.

Cost effectiveness of pollutant removal with regard to costs of construction, materials

and labour requirement for maintenance.

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