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BIOENERGY CROPS AND BIOREMEDIATION
BIOENERGY CROPS AND
BIOREMEDIATION - A REVIEW
A Contract Report by ADAS for the
Department for Food, Environment and Rural Affairs
Final Report
AUGUST 2002
BIOENERGY CROPS AND BIOREMEDIATION
CONTENTS
AUTHOR(S)
PAGE
EXECUTIVE SUMMARY
1.
INTRODUCTION
1
1.1
BACKGROUND
1
1.2
THE ENERGY CROPS IN
QUESTION
2
1.2.1
1.2.2
2
3
1.3
1.4
SRC willow and poplar
Miscanthus
4
4
4
5
THE ISSUES
1.3.1
1.3.2
1.3.3
2.
M Bullard & P Nixon
Land availability
Waste availability
The role of energy crops
6
REFERENCES
WASTE UTILISATION: ENVIRONMENTAL
& CROP EFFECTS
2.1
NITROGEN & PHOSPHORUS –
LOSSES TO WATER
2.1.1
2.1.2
2.1.3
2.1.4
2.1.5
2.2
2.2.3
P Johnson
Introduction
Nutrient requirements of
biomass crops
Use of organic manures or
wastes
Landfill leachate & other urban
aqueous wastes
References
GASEOUS LOSSES
2.2.1
2.2.2
7
Gaseous losses of nutrients
Emissions of radiatively active
gases from soils & organic
waste materials
References
7
7
7
11
14
16
J King
19
19
21
37
BIOENERGY CROPS AND BIOREMEDIATION
2.3
HEAVY METALS & OTHER
TOXINS
2.3.1
2.3.2
2.3.3
2.3.4
2.3.5
2.3.6
2.4
2.4.3
2.4.4
3.
F Nicholson
42
42
43
47
51
52
52
C Britt
Introduction
Agricultural and municipal
wastes
Industrial wastes
References
BIOREMEDIATION OF CONTAMINATED
SITES
56
56
60
66
68
P Nixon
73
3.1
INTRODUCTION
73
3.2
PHYTOREMEDIATION
73
3.3
POTENTIAL ENVIRONMENTAL
PROBLEMS
78
3.3.1
3.3.2
78
3.4
4.
PAGE
Introduction
Concentrations of
contaminants in livestock
manures & other wastes
Contaminant uptake/removal
by biomass crops
Other pathways of
contaminant movement and
environmental effects
Conclusions
References
IMPACTS ON BIODIVERSITY
2.4.1
2.4.2
AUTHOR(S)
Bioaccumulation of pollutants
Release of metals during
combustion
78
79
REFERENCES
WASTE UTILISATION AND
BIOREMEDIATION:
THE ROLE OF GMOs
4.1
INTRODUCTION
4.2
PHYTOREMEDIATION OF
INORGANIC POLLUTANTS
4.2.1
4.2.2
4.2.3
4.2.4
4.2.5
4.2.6
Root uptake
Transport within plants
Hyperaccumulation
Metallothioneins
Phytochelatins
Transformation of toxic
elements
N Smith
81
81
82
82
83
83
84
84
85
BIOENERGY CROPS AND BIOREMEDIATION
AUTHOR(S)
4.3
5.
PHYTOREMEDIATION OF
ORGANIC POLLUTANTS
N Smith
PAGE
86
4.4
PLANT TRANSFORMATION
88
4.5
RISKS
89
4.6
DISCUSSION
90
4.7
REFERENCES
91
LEGISLATION & CODES OF PRACTICE
AFFECTING RECYCLING & LAND
APPLICATION OF ORGANIC WASTES
5.1
INTRODUCTION
5.2
UK WASTE REGULATIONS &
DIRECTIVES
5.2.1
5.2.2
5.2.3
5.2.4
Codes of practice
Odours & gaseous emissions
Water pollution
Land application, including
bioremediation, of organic
materials
5.2.5 Draft Soil Strategy - England
5.2.6 Nitrate
5.2.7 Phosphate
5.2.8 Sewage sludge
5.2.9 Safe Sludge Matrix
5.2.10 Sewage sludge use in forestry
and on restored land
5.2.11 Waste Management Licensing
Regulations
5.3
G Hickman
98
98
98
100
101
102
103
103
104
105
106
106
108
109
UK LEGISLATION, NOT
PRIMARILY AIMED AT WASTE
DISPOSAL, WHICH MAY AFFECT
ORGANIC WASTES
110
5.3.1
5.3.2
5.3.3
5.3.4
110
111
111
111
Animal By-Products Order
Plant Health (GB) Order, 1993
Planning controls
Contaminated land
BIOENERGY CROPS AND BIOREMEDIATION
5.4
EC LEGISLATION: PROPOSALS
THAT MAY AFFECT ORGANIC
WASTE
5.4.1
5.4.2
5.4.3
5.4.4
5.5
6.
PAGE
G Hickman
112
112
Introduction
EC Sludge Directive – Working
Document 3rd Draft
EC Biological Treatment of
Waste Directive – Working
Document 2nd Draft
EC Landfill Directive
112
113
114
115
REFERENCES
CONCLUSIONS & RESEARCH
RECOMMENDATIONS
6.1
6.2
AUTHOR(S)
CONCLUSIONS
RESEARCH RECOMMENDATIONS
C Britt
118
118
119
BIOENERGY CROPS AND BIOREMEDIATION
EXECUTIVE SUMMARY
Introduction
1.
The Government's aim is for the UK to produce 10% of all electricity from renewable
sources by 2010. A viable biomass industry is central to this target. If energy crops
and forest residues are to provide fuel for an additional electricity capacity of 1,500
MW, around 125,000 ha of energy crops will be needed. Current plantings are of
willow (Salix) and poplar (Populus) short rotation coppice (SRC) and Miscanthus but,
although planting grants are available, uptake has so far been slow.
2.
Using energy crops for the disposal of agricultural, municipal or even industrial
wastes, or growing them on low-value ‘brown-field’ sites (e.g. capped landfill sites,
mining spoils and contaminated ex-industrial land), may provide additional revenue
and improve their appeal to growers. On brown-field sites energy crops also offer real
opportunities for site stabilisation and bioremediation of contaminated soils.
3.
This report summarises the findings of a desk study, which investigated the
opportunities and potential problems associated with systems that utilise energy crops
as disposal sites for waste materials or for the phytoremediation of contaminated land.
The study included an extensive review of the literature and expert analysis of the
issues. The main objectives of the study were:
a) To review opportunities for bioremediation with energy crops.
b) To evaluate the likely magnitude of available land and waste application rates.
c) To quantify possible nutrient and metal losses, via leaching in soils and water.
d) To quantify nutrient and metal uptake by energy crops.
e) To evaluate the potential for use of genetically modified organisms to improve
the bioremediation/waste utilisation capabilities of energy crops.
f) To evaluate the likely magnitude of atmospheric emissions of CH4, N2O and CO2
from wastes applied to energy crops.
g) To evaluate the impact of waste application, and/or siting on contaminated soils,
on biodiversity associated with energy crops.
h) To review and summarise the legal framework affecting disposal of farm and
non-farm wastes, and consider how relevant UK and EU legislation helps or
hinders the use of energy crops for bioremediation.
i) To identify research requirements.
Nitrogen and Phosphorus – Losses to Water
4.
Energy crops require little additional nutrients in the planting year. In particular, extra
nitrogen is not required, as mineralisation from soil organic matter will supply
sufficient quantities of this nutrient and applying more will increase nitrate losses by
leaching. Once established, atmospheric deposition and nutrient return in leaf fall are
likely to be significant factors in energy crop nutrient cycles.
5.
There is conflicting evidence on the N requirements of energy crops from the second
year onwards, possibly arising from the dynamics of nutrients within the crop. Dense
SRC crops have a higher proportion of bark in the harvested wood, and bark contains a
BIOENERGY CROPS AND BIOREMEDIATION
much higher concentration of nutrients. In SRC and Miscanthus, Danish work
suggests leaching does not occur, even when organic manures are applied at rates well
above standard recommendations. However, these could be mineralised, and leached
from the soil at harvest, or after restoration of the site to agriculture. There is also the
likelihood of large CO2 losses from the soil following site restoration.
6.
To avoid nutrients leaching into surface or ground-water, only manures or slurries with
a low available N content should be applied prior to planting energy crops, or in the
establishment year. Organic wastes with higher available N contents might be applied
from the second year onwards.
Applications should be guided by the
DEFRA/NAWAD Code of Good Agricultural Practice for the Protection of Water
(the ‘Water Code’), and take into account crop off-take.
7.
The Water Code specifies that a maximum of 250 kg ha-1 of organic nitrogen can be
applied in any twelve month period. This figure is used throughout this review, for
illustrative purposes. It should be noted that this figure is for total organic nitrogen
and that available nitrogen levels will depend on the type of organic manure/waste
being applied. For low available nitrogen materials, in non-sensitive catchments, 500
kg N ha-1 is allowed every two years. With the expansion of nitrate vulnerable zones
(NVZs) in 2002 at least 55% of England will fall within sensitive catchments.
8.
Regular applications of sewage sludge or manures are likely to rapidly increase soil
phosphate levels and the potential for phosphate losses by leaching. Energy crops
have a relatively low demand for phosphate and should not be used for the continued
disposal of high phosphate organic wastes.
9.
Irrigating crops with the leachate from buried waste requires further research.
Gaseous Losses
10. Increased gaseous flux of nitrogen and sulphur may occur as a result of the application
of organic wastes to energy crops. Ammonia volatilisation from manures and slurries
inevitably occurs and, typically, 40-50% of the plant available N can be lost within six
hours of application. Whilst the most effective way of reducing ammonia
volatilisation from slurries is to inject the slurry into the soil, this is not an option in
energy crops, due to the mass of roots or rhizomes. Application methods and rates of
manure use that greatly reduce soil air-space or gaseous exchange will increase
emissions of nitrous oxide (N2O) and methane (CH4) ‘greenhouse gases’.
11. Manure applications are likely to closely follow crop harvests, exacerbating soil
compaction caused by harvesting machinery. These factors suggest that losses of N2O
from energy crops following organic waste application are likely to be high but losses
have not yet been quantified. Emissions of N as N2O from sewage sludge, over a full
year, are likely to be of the order of 1-2% of applied N – similar to those from
livestock wastes.
12. Production of energy crops will result in net increases in soil organic matter (through
microbial decomposition of the leaf litter and turnover of fine roots) and below-ground
carbon, stored in a semi-permanent root system. This may be enhanced by regular
applications of organic wastes.
BIOENERGY CROPS AND BIOREMEDIATION
13. Using existing figures, the net effects of applying organic wastes to energy crops on
the contribution to radiatively active gases in the atmosphere and carbon sequestration
were estimated. If livestock wastes are applied to SRC every three years, after harvest,
at the maximum rates allowed by the DEFRA Water Code, the estimated benefits of
carbon sequestration may be reduced by the effects of CH4 and N2O emissions by
between 27 and 36%. The corresponding figures estimated for Miscanthus, which
sequesters slightly more carbon, suggest a 24-31% reduction in C sequestration. If
wastes were applied annually to Miscanthus nearly all of the carbon sequestration
benefits would be negated by N2O (mainly) and CH4 emissions, by approximately 6875%. These estimates are based on sequestration/emissions over a full economic
rotation of 25 years.
14. With the above provisos, the applications of moderate quantities of organic wastes to
energy crops is likely to be beneficial to soil sustainability and to make modest
contributions of carbon sequestration. As well as providing an important source of
plant nutrients and valuable quantities of organic matter, livestock manures and other
organic wastes applied to fuel crops may contain compounds that can be harmful when
applied to plant ecosystems in the human food chain.
Heavy Metals and Other Contaminants
15. Heavy metals [including arsenic (As), cadmium (Cd), chromium (Cr), copper (Cu),
lead (Pb), mercury (Hg), nickel (Ni) and zinc (Zn)] cannot be broken down into less
harmful by-products, so phytoremediation strategies focus on their accumulation in
above-ground plant parts and removal from the contaminated site. Uptake is affected
by the type of waste and the soil type.
16. Organic pollutants can often be completely broken down by plants into less harmful
metabolites. Research on hybrid poplars has demonstrated their ability to take up and
effectively degrade or deactivate a number of other contaminants, including atrazine,
1,4-dioxane, TNT and trichloroethylene.
17. Biosolids, rather than livestock manures, will be a more important source of heavy
metals and organic contaminants to biomass crops. Although research evidence is
limited, the current consensus is that organic contaminants in biosolids applied to
agricultural land are unlikely to cause significant environmental or human health
problems. However, whilst the risk of soil metal contamination from biosolids is
recognised and regulated for, there is no current control over inputs of heavy metals
from livestock manures, and care should be taken to ensure that soil concentrations do
not become excessively high, especially where pig and poultry manures are applied.
18. High yielding biomass crops offer good potential for the phytoremediation of sites
contaminated with heavy metals. Willows have been shown to take up large amounts
of Cd and Zn. Different biomass crops, species and clones may show large differences
in efficiency of heavy metal uptake. There can also be large differences in the
concentration of metals in different plant parts. For phytoremediation purposes, it is
desirable for metals to be concentrated in the harvested parts of the plant e.g. SRC
stems/wood.
19. The maximum annual applications of high-metal content organic wastes, such as
sewage sludge or poultry manure, to willow SRC crops are likely to lead to a net
BIOENERGY CROPS AND BIOREMEDIATION
accumulation of metals (e.g. Zn, Cu, Pb, Ni and Cr) in the soil, as application rates
would exceed crop uptake.
However, on average soils the maximum soil
-1
concentration of 200 mg Zn kg would be reached in around 110 years for sludge and
160 years for poultry manure.
20. Metals removed from sites in energy crops will primarily be concentrated in the ash
after combustion or gasification, except mercury which may volatilise and be emitted
from the stack to be deposited onto soils in the vicinity of the combustion plant. Other
metals may be discharged into the atmosphere, via particulates in the flue gases, and
subsequently re-deposited onto land.
21. Recovery of metals from ash may be technically possible, but very expensive. The
disposal of ash with high heavy metal concentrations may pose problems.
Impacts on Biodiversity
22. Research has shown the value of willow and poplar SRC as a habitat for certain animal
groups, including phytophagous insects, songbirds (especially warblers) and
pheasants. There has been very little research on the ecology of Miscanthus or other
energy grasses.
23. The extent and composition of the ground flora will, to a large degree, determine the
biodiversity of animal species within the crop. In most SRC plantations the ground
flora is either sparse, due to effective weed control measures, or of limited diversity
and dominated by species of low conservation value, typical of agricultural weed
communities or disturbed land. Species such as bramble may invade coppice later in
the cycle. One extensive survey of SRC plantations showed common nettle to be the
most frequent species (81% of all sites).
24. Animal groups known to be well represented in SRC plantations include herbivorous
invertebrates that feed on the crop foliage, like willow beetles, and their predators and
parasites; insectivorous woodland birds (e.g. thrushes, tits and warblers); and
mammals such as rabbits, roe deer and wood mice. Populations of ground-dwelling
species will be heavily dependent upon the extent and species composition of the
ground flora.
25. The application of waste materials might be expected to have significant effects on the
flora and fauna of energy crops. However, no evidence was found in the literature of
research that directly addressed this issue. There was only a limited amount of
published research on the ecological effects of applying relevant agricultural, urban or
industrial wastes to other types of vegetation.
26. Thick applications of waste materials, particularly those that are slow to decompose,
may have a ‘mulching’ effect and suppress the ground flora, whilst organic wastes
may provide a valuable additional food source for soil and ground-dwelling microorganisms and invertebrates. Mulching would have a net detrimental effect, whilst
provision of a supplementary food source would be generally positive. Both would
have ‘knock-on’ effects up through the food chain.
27. The addition of nutrients through organic waste applications will also have indirect
effects on the growth and composition of the ground flora, and provide crop nutrients.
BIOENERGY CROPS AND BIOREMEDIATION
However the potential ecological effects of nutrient inputs from manures, slurries and
sludges can probably be regarded as minimal, as similar quantities of nutrients would
otherwise be applied in inorganic fertilisers. What may be more significant are the
effects of additional water supplied in slurries or dirty water.
28. The bioaccumulation of heavy metals or organic toxins in animal tissues, following the
application of contaminated wastes to energy crops, is an important concern.
However, although high levels of heavy metals (e.g. Cd) in the livers and kidneys of
insectivorous shrews and deer species have been observed in US studies, generally the
heavy metal contents of these mammalian organs were not considered to have
important effects on the health of the animals concerned. It is not known what the
effects of low concentrations of compounds like polycyclic aromatic hydrocarbons
(PAHs) in animal manures and slurries might have on soil-inhabiting invertebrate
groups in biomass crops.
29. At the top of the food chain, predators such as foxes, stoats and birds of prey, might
appear to be at greatest risk, but the relatively large ranges of most predator species
lessens the degree of exposure to contaminated prey. The risks would, however,
greatly increase in the event of a significant area of waste-treated biomass crop being
situated in a single location. Sewage contaminated with pathogens pose another risk to
wildlife, and although, in most cases, risks may be low, research has shown that in
some situations deer may preferentially graze areas that have been treated with sewage
sludge, thus increasing their exposure risk.
Bioremediation of Contaminated Sites
29. The necessity to decontaminate polluted sites is recognised, both socially and
politically, because of the increasing importance placed on environmental protection
and human health. As the number of sites and levels of contamination rise, so does the
need to develop effective and affordable methods for decontamination.
30. Phytoremediation is the term used to describe the use of plants to mitigate the effects
of contamination.
There are four fundamental processes that make up
phytoremediation: phyto-immobilisatio, phyto-stabilisation, phyto-extraction and
phyto-volatilisation. Phytoremediation is a low-cost option, particularly suited to large
sites that have relatively low levels of contamination.
31. Many species of Salix, Populus and Miscanthus have characteristics of ‘pioneer’
species – with adaptations for growth on poor sites, under harsh conditions. Willow
varieties differ significantly in their levels of tolerance and rates of uptake of heavy
metals (see Exececutive Summary 15).
32. Results from pot studies indicate that Miscanthus crops could be successfully grown
on contaminated land, although high levels of heavy metals may reduce crop
productivity. Most heavy metals accumulate in the roots and rhizomes, rather than in
the harvested aerial parts.
BIOENERGY CROPS AND BIOREMEDIATION
Waste Utilisation and Bioremediation: the Role of Genetically Modified
Organisms
33. The genetic modification of biomass crops to improve pollutant uptake, transport,
accumulation and tolerance, offers the potential to dramatically increase the
effectiveness of phytoremediation of organic compounds and metals from
contaminated sites. A single GM energy crop might be produced to efficiently take up
several different pollutants.
34. Poplar species have been the subject of a considerable amount of research on the
introduction and over-expression of foreign genes. The first reports of poplar
transformation were published in the late 1980s. The potential genetic transformation
of willows and Miscanthus has received relatively little attention. However, for
Miscanthus, research on embryogenic suspension culture, and other in vitro
propagation systems, and the transformation of callus tissue via microprojectile
bombardment, has progressed to the point that the generation of transgenic Miscanthus
clones can be expected in the relatively near future.
35. Genetically modified biomass crops offer the potential for dramatically improved
phytoremediation capabilities, but careful consideration must be given to the
environmental consequences before such crops are exploited. Primary considerations
should include the potential for bioaccumulation of toxic pollutants in animal
populations (a problem that is likely to be exacerbated by hyperaccumulation of
pollutants in the stems and leaves of GM plants) and the prevention of any possible
gene flow into ‘conventional’ crop plants/native plant populations.
Legislation and Codes of Practice Affecting the Recycling and Land
Application of Organic Wastes
36. The use of wastes, sludges and manures on land generally, and on biomass crops is
covered by UK legislation and codes of practice, and EC directives. The regulations
are continually changing as old legislation is updated and new controls and
recommendations are issued, that must be considered before organic waste materials
are applied to land.
37. The DEFRA/NAWAD Codes of Good Agricultural Practice for the Protection of
Water, Air and Soil (1998) are statutory codes under Section 97 of the Water
Resources Act 1991 and give guidance on best practice for avoiding pollution. Most
farmers and other operators applying organic wastes to land follow the guidance of the
Codes and all Water Operators have signed up to compliance with them. Compliance
is also a basic requirement for participation in most Farm Assurance Schemes. As
with all Codes of Practice failure to comply with the Codes is not an offence, but
would be taken into account in any legal action taken as a result of a pollution
incident.
38. Under the Pollution Prevention and Control (England & Wales) Regulations 2000,
industrial installations are regulated by the Environment Agency or local authorities.
The impacts of these relatively new regulations are still hard to assess, but they are not
expected to hinder the use of wastes on energy crops. The recycling of nutrients in
livestock manures is excluded from EC and UK regulations on the use of ‘waste
materials’ spread to land.
BIOENERGY CROPS AND BIOREMEDIATION
39. Legislation covering water catchments, and the earlier compliance measures to meet
the requirements of EC Nitrate Directive (91/676/EEC) are reviewed and the effects of
compliance commented on. The recently announced expansion of the areas covered
by Nitrate Vulnerable Zones make the restrictions far more widely applied, but the
ability of growers of biomass crops to comply remains much the same; the Action
Programme limits to nitrogen use are those in the Codes of Practice.
40. The issue of phosphate enrichment of water is becoming increasingly important. The
Soil and Water Codes provide guidance on limiting phosphate losses and recommend
that farmers should avoid total P inputs in excess of crop off-take. The apparently low
P requirements of bioenergy crops could limit the total amount of organic wastes that
can be applied.
41. Sewage sludge applications to land, and biomass crops, are regulated by EC Directive
86/278/EEC (1986), and should comply with the Water Code, with applications
supplying no more than 250 kg total N ha-1 yr-1. For farmyard manure, the Water
Code allows an option for low N sewage products, such as sludge cake, to be applied
at 500 kg total N ha-1 in every other year (in non-sensitive catchments only – see
Executive Summary 6). Revised regulations will adopt the provisions of the Safe
Sludge Matrix, and introduce standards for pathogen content in sludges. They will
also ban the use of untreated sewage sludge on all crops from 31 December 2005
(They have been banned on all food crops since December 1999).
42. The use of sewage sludge on non-agricultural land, and for land restoration, is
controlled by the Waste Management Licensing Regulations 1994. These regulations
do not, however, cover wastes arising from agriculture. Also excluded are numerous
specified materials for use in agriculture. Disposal of exempt materials must, however,
result in agricultural or ecological benefits, and are generally subject to a maximum
annual application rate of 250 t ha-1.
43. The Town and Country Planning Act 1990 (as amended by the Planning and
Compensation Act 1991) may limit the use of certain wastes on energy crops that
might be planted after mineral extraction has ceased, whilst requirements of the
Environmental Protection Act 1990 and Environment Act 1995 may act as drivers for
the planting of biomass crops for the bioremediation of contaminated land identified
by local authorities as potential risks to human health or the environment.
44. A new proposed EC Sludge Directive may severely restrict the recycling of sewage
sludge, and certain other organic materials to agricultural land. It is likely to
significantly reduce the maximum levels for heavy metals in soils, introduce limits on
levels of heavy metals and organic compounds in sludges, and include treatment
standards based on pathogen reduction. It is also likely to confirm the principle that
mixing any sludge with other non-regulated wastes makes the entire combined waste
stream subject to the conditions of the Directive.
45. Similarly the proposed EC Biological Treatment of Biowaste Directive (or ‘Compost
Directive’) would severely restrict the treatment and application to land of many
mixed wastes and industrial wastes. The proposed maximum application rates for
organic wastes are generally much lower than are currently applied in the UK. Both
the Sludge and Compost Directives are unlikely to be enacted in UK legislation before
2005.
BIOENERGY CROPS AND BIOREMEDIATION
46. The EC Landfill Directive 1993/31 requires significant reductions in the quantities of
putrescible materials disposed of in landfill sites. Targets have been set for 2006,
2009 and 2016. An estimated five million tonnes of organic material is disposed of
annually within the UK household waste stream. The Directive is driving waste
management companies and waste producers to look at land spreading and
bioremediation, as a possible route to divert this material away from landfill.
BIOENERGY CROPS AND BIOREMEDIATION
CHAPTER 1
INTRODUCTION
MIKE BULLARD & PETER NIXON
1.1
BACKGROUND
In response to now incontrovertible evidence of human-induced global warming
due to the emission of atmospheric warming gases (IPCC, 2000), regional, national
and European targets for both emissions reduction and renewable energy generation
have been set (DTI, 1999; EC, 1998). For the UK, a target of 10% of electricity
generation from renewables has been set for 2010. Central to the government’s
renewable energy targets is the stimulation of viable biomass crops (DTI, 1999).
It is anticipated that 1,500 MW of new electrical capacity might come from energy
crops and forestry residue combustion, and it has been proposed that 125,000 ha of
energy crop planting might be needed in order to meet this target.
In the UK, the energy crops most likely to be grown are short rotation coppice
(SRC) willow and poplar and Miscanthus grass. These high-yielding species are
eligible for establishment grants of up to £1,600 ha-1 on IACS-registered arable or
grassland, or other grazed grassland (see MAFF, 2000 for full details). Sufficient
grants are currently available to enable 23,000 ha of crop to be established; yet less
that 2,000 ha have been planted in England and Wales (At the time of writing all of
this is willow coppice, mainly established by ‘Project Arbre’ in Yorkshire,
Nottinghamshire and Lincolnshire).
Major disincentives for uptake of the crops are:
1.
Energy generation schemes (the ‘end user’) are coming on-stream at a much
slower rate than anticipated, thus demand for the crops is lower than
anticipated.
2.
Establishment grants only cover 40% or 50% of establishment costs (for
Miscanthus and SRC, respectively), leaving a deficit of at least £1,000 ha-1.
3.
Energy crops require a long-term commitment for relatively low financial
reward. Other less risky cropping options are, therefore, more attractive.
4.
Yields from the first plantings have been disappointing and this has
suppressed interest.
For these reasons, crop uptake has been low. In order to encourage energy crop
planting it is, therefore, necessary to consider other initiatives where additional
economic benefits can accrue from growing the energy crop.
Increasing interest is being given to the concept of disposal of agricultural and
municipal wastes on energy crops. This potentially provides organic matter and
nutrients needed for crop growth at a low cost, whilst enabling controlled disposal
of wastes on a non-food crop. Indeed, some grower organisations recommend the
application of wastes to provide the necessary nutrients for SRC production.
1
BIOENERGY CROPS AND BIOREMEDIATION
Alternatively, it is pertinent to consider biomass crop opportunities on nonagricultural land, where growing the crop may provide an additional benefit to the
land owner. For example, the biomass crop may provide essential site stabilisation
(and, possibly, decontamination) - with the sale of biomass an additional, but nonessential, benefit. Many are now proposing that energy crops should be grown on
brown-field sites (e.g. capped landfill sites, quarries, mine spoils or contaminated
ex-industrial land), in order to facilitate bioremediation of heavy metals and toxins.
Another advantage here is that land prices are much lower than for agricultural
land, and energy crops are a true added benefit to a situation where the major
priority is decontamination of the land.
Whilst both routes offer advantages, the true extent of the advantages and the
magnitude of the land availability for such cropping are poorly understood.
Equally, there are also a number of potential hazards from using these disposal
routes and/or sites. Whilst there is information in the public domain about specific
examples of waste disposal or bioremediation, there is no consolidation of the
subject in a way that would allow policy makers to make clear judgements on the
needs and opportunities posed. For DEFRA and Government to support such
energy cropping initiatives, there must be clear information on the magnitude of the
risks and opportunities that such disposal routes present.
This review brings together detailed information on these issues and also identifies
areas requiring further research.
1.2
THE ENERGY CROPS IN QUESTION
Any crop or agricultural residue that can be presented in a relatively dry form (i.e.
at approximately 25% moisture content, or less) is suitable for thermo-chemical
conversion. Material with high moisture content can be efficiently used in
combined heat and power (CHP) plants. The purpose of growing energy crops is to
maximise the yield of ligno-cellulose that can be harvested and presented to a
power station in a useable form. The two crops that offer this yield advantage are
now described in more detail:
1.2.1 SRC willow and poplar
Short rotation coppice (SRC) consists of densely planted, high-yielding varieties of
willow, and occasionally poplar, harvested on a two to five year cycle, although
most commonly every three years. SRC is a woody, perennial crop, the rootstock
or stools remaining in the ground after harvest with new shoots emerging the
following spring. An SRC plantation should be viable for up to 30 years before replanting becomes necessary.
The osier (Salix viminalis), a native shrub or small tree, is the parent species of the
majority of willow varieties grown as energy crops.
Willow SRC is mechanically planted in the spring. Unrooted hardwood cuttings,
produced by specialist breeders, are inserted into cultivated soil, using equipment
2
BIOENERGY CROPS AND BIOREMEDIATION
specifically designed for fast and efficient planting. The planted material grows
rapidly in the first year, reaching up to 4 m in height - depending on soil conditions.
During the first winter after planting, stems are cut back to ground level to
encourage the growth of multiple stems, that is the true coppice. Generally three
years after cutback, and again during the winter, the crop is harvested. The
equipment used for harvesting will depend on the requirements of the
customer/end-user, that is their fuel specification, but in all cases the harvesting
equipment will have been specifically developed or have involved modification of
existing harvesting machinery. The majority of other operations - such as land
preparation, spraying and fertilising - can be completed utilising conventional farm
machinery.
In the UK, yields achievable from willow SRC at first harvest are normally
expected to be between 7 and 12 oven dry tonnes (odt) per hectare per year (odt ha-1
yr-1), depending on ground conditions and efficiency of establishment. Yields
should increase at second and third harvests, as the stools mature. Plant breeding
programmes are continuing to identify further new willow varieties, that will
produce higher yields and demonstrate improved resistance to pests and diseases.
1.2.2 Miscanthus
Miscanthus species are woody, perennial, rhizomatous grasses, originating from
Asia, which have the potential for very high rates of growth. Miscanthus may be
familiar to many as a flowering garden ornamental, but it is the non-flowering
forms that are of interest agriculturally.
Miscanthus is spring-planted, and canes produced during the summer are harvested
in winter. This growth pattern is repeated every year for the lifetime of the crop,
which will be at least 15 years.
Miscanthus differs from short rotation coppice willow in that it gives an annual
harvest, and thus an annual income to the farmer. Miscanthus spreads naturally by
means of underground storage organs (rhizomes). However, their spread is slow
and the risk of uncontrolled invasions of hedgerows or fields is, consequently,
considered to be very low.
Rhizomes can be split and the pieces re-planted to produce new plants. All
propagation, maintenance and harvest operations can be done with conventional
farm machinery.
In the UK, long-term average harvestable yields from a mature crop have exceeded
13 odt ha-1 yr-1 at the most productive experimental sites. These high yields suggest
that the crop has the potential to make an important contribution to the UK’s
commitments to energy generation from renewable sources.
3
BIOENERGY CROPS AND BIOREMEDIATION
1.3
THE ISSUES
The disposal of agricultural and municipal wastes to land and the use of
contaminated sites for non-food cropping are two areas of increased concern for
policy-makers.
1.3.1 Land availability
The Environment Agency has estimated that there are between 50,000 and 300,000
ha of contaminated land across the country – distributed between some 100,000
sites. Of these sites, between 5,000 and 20,000 require remedial action, as specified
under legislation.
In the absence of any formal surveys it is difficult to provide accurate figures on the
extent of contaminated land in each region. However, ‘derelict land’ - defined as
'land so damaged by industrial or other development as to be incapable of
beneficial use without treatment', occupies an estimated 39,000 ha of land across
England (DoE, 1993).
Mines and quarries represent a significant area. A 1994 survey of land for mineral
workings in England indicated that just under 14,500 ha of land had planning
permission for surface disposal of mineral working deposits. Nearly 9,900 ha were
actually affected by tipping (although some sites were in the process of being
reclaimed), and over 6,000 ha had planning conditions for reclamation.
Thus, as much as 1-1.5% of the UK land area may be considered derelict or in need
of remediation, and much of this may be potentially available for energy cropping.
1.3.2 Waste availability
As can be seen from Table 1.1, significant quantities of wastes are produced each
year from various land-based sectors in the UK (DEFRA, 2001).
Table 1.1. Estimated total annual waste production in the UK, by sector
Sector
Annual raisings (Mt)
% of total
86
116
16
43
30
26
51
73
428
20
27
<1 6
10
7
6
12
17
100
Agriculture 3
Mining and quarrying 1
Sewage sludge 2
Dredged spoils 1
Municipal 4
Commercial 2
Industrial 2
Demolition and construction 5
TOTAL
1
2
Estimates for 1997
Estimates for 1998/99
3
4
Estimates for 1999
Estimates for 1999/2000
4
5
6
Estimates for 2000
Figures based on estimated dry wt
(26.5 Mt total wet wt.)
BIOENERGY CROPS AND BIOREMEDIATION
1.3.3 The role of energy crops
Energy crops have the potential to utilise agricultural and municipal wastes, and to
stabilise or clean up contaminated land. From the perspective of waste
disposal/utilisation, energy crops offer the following potential benefits:

they are not going to enter the human food chain;

they are perennial crops, thus allowing long-term breakdown of organic
matter in soils prior to renovation to food cropping;

they produce large quantities of biomass that, theoretically, requires large
quantities of nutrients, and thus are a sink for the nutrients in waste.
From the perspective of bioremediation of contaminated sites, they offer the
following potential solutions:

they utilise land that would otherwise have no agricultural value;

they are non-food crops that will not enter the human food chain;

they are perennial crops which may act as ‘excluders’ of contaminants in the
soil;

alternatively, they may act as ‘tolerators’ of the contaminants, actively
taking up the elements which, in some instances, can then be recovered
during biomass combustion;

the crops can also act as bioremediators of liquid leachates produced from
rainfall onto landfill and other contaminated sites;

in these situations, they may also act as recipients of agricultural and
municipal wastes.
There are attendant risks with such systems. With respect to the application of
agricultural and municipal wastes, these include:

risks of leaching nutrients applied in sludges into groundwater;

risks of increased atmospheric emissions of ‘greenhouse’ gases, associated
with global warming. For example, sewage sludge can increase the
emission of nitrous oxide (Scott et al., 1998) and maybe methane, which
although emitted in trace amounts are more effective greenhouse gases
(CH4 = 32 CO2 equivalents; N2O = 150 CO2 equivalents; Bouwman,
1990);

risks of contaminant accumulation in the production system, which are then
emitted from power station stacks upon combustion of the biomass;

negative impacts on the biodiversity associated with energy crops.
With respect to bioremediation, the risks include:

risks of contaminant accumulation in the production system, which are then
emitted from power station stacks upon combustion of the biomass;

negative impacts on the biodiversity associated with energy crops.
The practices, therefore, have the potential of negating some benefits derived from
saving fossil fuels by growing energy crops. There is a current lack of evidence on
the thresholds of application that are acceptable in different situations and this is, in
part, due to a lack of primary research in many areas.
5
BIOENERGY CROPS AND BIOREMEDIATION
In order to assist DEFRA and other stakeholders in the production of such research
priorities the following review condenses the research that has been carried out to
date on these issues, and draws parallels with other cropping systems.
1.4
REFERENCES
Bouwman, A F (Ed.) (1990). Soils and the Greenhouse Effect. John Wiley and
Sons, Chichester. 575 pp.
DEFRA (2001). The Environment in Your Pocket 2001. http://www.defra.gov.uk/
environment/statistics/eiyp. Department for Environment, Food and Rural
Affairs.
DoE (1993). Survey of Derelict Land in England and Wales. Department of
Environment.
DTI (1999). New and Renewable Energy: Prospects in the UK for the 21st
Century. HMSO, London.
DTI (1999). Waste Strategy 2000 for England and Wales: Parts 1 & 2 - ISBN 0
10 146932 2 / 0 10 146933 0. HMSO, London.
EC (1998). Energy for the Future: Renewable Sources of Energy. European
Commission White Paper.
IPCC (2001). Third Assessment Report – Climate Change 2001:The Scientific
Basis. Special report on emissions scenarios (Eds. J Houghton, Y Ding, D J
Griggs, M Noguer, P J van der Linden, X Dai, K Maskell & CA Johnson).
Intergovernmental Panel on Climate Change, United Nations.
MAFF (2000).
England Rural Development Plan: Energy Crop Scheme
Explanatory Booklet. HMSO, London. 24pp.
Scott, A; Ball, B C; Crichton, I J & Aitken, M N (1998). Nitrous oxide
emissions from grassland amended with sewage sludge. Short communication
in Soil use and Management, 14. 55.
6
BIOENERGY CROPS AND BIOREMEDIATION
CHAPTER 2
WASTE UTILISATION: ENVIRONMENTAL AND CROP EFFECTS
2.1
NITROGEN AND PHOSPHORUS – LOSSES TO WATER
PADDY JOHNSON
2.1.1 Introduction
Consideration of the potential benefits or disadvantages of applying novel waste
materials to short rotation coppice or graminaceous biomass crops should be
preceded by a philosophical assessment. Are the materials being applied to
improve crop growth, as a method of waste disposal or in order to reduce adverse
environmental effects? Indeed will the use of biomass crops for cleaning up wastes,
for example exploitation of the ability of Salix to remove cadmium, increase other
adverse environmental effects? Possibly this is beyond the scope of this study.
This section of the report, therefore, starts with a review of the nutrient
requirements of biomass crops, so that the potential for increasing potential losses
of nutrients to the environment can be assessed. It is unfortunate that there appears
to be little published information on the beneficial (or otherwise) effects of the
application of many of the wastes which it is being suggested could be applied to
biomass crops, and none on the consequences of removing the biomass crop once
its useful life has ended. It is important that long term effects are discovered, for
example if biomass crops are used for the disposal of sewage sludge, will there be a
large loss of nitrogen and carbon dioxide to the environment when the plantation is
removed and the land returned to conventional agriculture?
2.1.2 Nutrient requirements of biomass crops
The nitrogen cycle in the soil is complex. Mineralisation of N from soil organic
matter plus crop residues, together with inputs from the atmosphere, can supply
large quantities of available N. Mineralisation is likely to be particularly high in the
year of planting following disturbance of the soil (Shepherd et al., 1996), and
decrease thereafter - unless weed control is by mechanical means. Applied or
mineralised nitrogen can be lost from the soil by leaching or by denitrification in
wet warm conditions. A balance sheet approach to N fertilisation has some
attractions, but many of the factors to be taken into account are not well understood.
Several authors have indicated that no nitrogen is required in the planting year for
short rotation coppice (Ford-Robertson et al., 1991; Kopp et al., 1993; Ledin &
Alriksson, 1996; McLaughin et al., 1985). Indeed, application of soluble fertiliser
at planting could have a deleterious effect on growth, as it raises the osmotic
potential of the soil solution and may thus slow root development. Few countries
give a recommendation for some nitrogen at this stage (Table 2.1); in subsequent
years recommendations vary.
7
BIOENERGY CROPS AND BIOREMEDIATION
Table 2.1. Nitrogen recommendations from various countries
Source
Recommendation in
establishment year
Subsequent years
45
100-150
0
?
0
0
?
60
122-217
50?
100-135
112-168
Ledin & Alriksson (1996),
Sweden
Chambers & Mitchell (1995),
UK
van Veen (1981), Netherlands
Frison (1995), Italy
Canada
USA
Note:
Data from Ledin & Willebrand (1996) unless otherwise indicated.
Similarly there are suggestions that Miscanthus does not respond to the application
of nitrogen (SORGHAL, 1997).
Nitrogen requirements post-planting
Experience in the United States, summarised by Christopherson (1996), suggests
that, in poplar plantations on sites with moderate or good levels of fertility and good
weed control, N is unlikely to be required for at least two years. Furthermore,
Christopherson (1996) suggests that, due to the recycling of nitrogen in leaf fall,
applications to poplar plantations after canopy closure can be greatly reduced or
even omitted on some sites.
On sandy loam and loamy sand soils in Denmark, Mortensen et al., (1998) obtained
a significant increase in yield where N was applied at 75 kg ha-1 in the second year
on one site, but there was a non-significant reduction in yield in the third year
(Table 2.2). Despite low mineral nitrogen content of the soils, responses at these
sites may have been restricted by low water availability.
McLaughlin (1985) needed to wait until the third year before recording any
increase in yield of poplar wood related to the application of nitrogen. In Northern
Ireland, Dawson (1999) has suggested that there was no response to the application
of N for at least two rotations when the coppice was grown on reasonably fertile exarable land.
Table 2.2. Response of SRC to nitrogen applications
Site
Soil
Type
Foulam
Jyndevad
*
loamy
sand
sandy
Year 1
Year 2
Year 3
0N
75 kg ha-1 N
0N
75 kg ha-1 N
0N
75 kg ha-1 N
1.25
1.32
2.68*
3.71*
8.71
6.93
1.62
1.64
3.52
4.94
3.67
3.61
Statistically significant response (P < 0.05)
8
BIOENERGY CROPS AND BIOREMEDIATION
Christopherson (1996) suggested that a leaf analysis level of 3% N in the dry matter
in mid-July, indicates a sufficiency of nitrogen in the system. This is a leaf
concentration slightly above that used for top fruit in the UK (MAFF, 2000). Leaf
analysis could be the best way of assessing nitrogen need and should be
investigated.
Hall et al. (1996) recommended that growers of short rotation coppice should be
discouraged from applying nitrogen fertiliser.
It is possible that Miscanthus does not respond to nitrogen applications during post
establishment years (SORGHAL, 1997).
Other nutrients
There is a paucity of reports on the response of biomass crops to nutrients other
than nitrogen. Therefore a balance sheet approach for the maintenance of soil
reserves may be the best approach for deriving crop requirements. An exception is
a report by Frison (1995), which differentiates between tree populations, based on
the fact that the majority of the nutrients are in the bark and thus the percentage of
bark will be much higher at the higher populations (Table 2.3). In the early years of
growth uptake and potential removal of phosphate and potash will be small. Of
these two nutrients only phosphate carries an environmental threat. Except at very
low soil levels applications of these nutrients based on a balance sheet approach
would seem appropriate.
Table 2.3. Nutrient off-takes in timber of Populus x canadensis clone I-214,
kg per oven-dry tonne (Frison, 1995)
Nitrogen
Phosphate
Potash
75,000 trees ha-1
6.9
3.1
5.2
10,000 trees ha-1
5.6
2.1
4.2
-1
1.6
0.7
2.4
277 trees ha
Losses of nitrogen by leaching
Nitrogen applied at planting or early in the life of the crop is readily lost from the
soil by leaching. Mortensen et al., (1998) reported that where 75 kg ha-1 N was
applied at planting, leaching losses of N were 32 kg ha-1 higher than where no N
was applied. Christian & Riche (1998) showed large leaching losses of N under
establishing Miscanthus, which were further increased by the application of N
fertiliser on a site already rich in mineralised N.
On a moisture retentive soil in west Norfolk, poplars were grown at 3.5 m spacings
and fertilised with ammonium nitrate (60, 180 and 300 kg ha-1) or pig slurry applied
in spring to supply similar amounts of nitrogen. Where the higher rates of N were
9
BIOENERGY CROPS AND BIOREMEDIATION
applied, both tree height and incremental diameter improved dramatically.
However, there was a downside, which could be partially controlled by other
management techniques. Nitrate leaching was significantly increased (Johnson,
1995) at the higher nitrogen rates (Table 2.4). The experiment showed that where
manures are applied, environmental effects needed to be taken into account. If
manures were applied for several years and nitrate losses controlled by husbandry
(grass would not be the preferred option as it will reduce the amount of moisture
available for tree growth), there was a potential for a large build-up of soil N to
occur. This might be mineralised when the soil is disturbed at harvest or the
plantation removed.
Table 2.4. Nitrate-N losses (kg ha-1) by leaching at ADAS Terrington,
Norfolk over two winters
N applied in slurry (kg ha-1)
Weed control
Ammonium nitrate
0
60
180
300
300
Cultivation
86
83
133
265
212
Herbicide
48
17
32
110
200
Grass strip
3
5
2
24
39
Weed control
0.002
17.4
4
P
SE
df
N rate
< 0.0001
36.6
24
Interaction
0.002
36.6
24
Further monitoring of this site (Johnson, 2000) during winter 1999/2000, after the
original treatments had been abandoned some five years earlier, showed much
reduced levels of nitrogen loss (Table 2.5), although drainage was very low.
Table 2.5. N loss and mean N concentration in winter 1999/2000, estimated
from porous ceramic cups installed at 90 cm depth
Cult
Nil N
Grass
NH4NO3
Cult
Grass
Pig slurry
Cult
Grass
Fertiliser
SEM Prob.
Cover
SEM Prob.
N loss
(kg ha-1)
0.27
0.08
0.38
0.05
0.99
0.16
0.08
0.01
**
0.03
0.02
**
Mean N
conc.
(mg l-1)
1.31
0.39
1.81
0.24
4.74
0.79
0.37
0.01
**
0.13
0.02
**
**
Statistically significant response (P < 0.01)
Note:
Based on drainage of 21 mm.
10
BIOENERGY CROPS AND BIOREMEDIATION
Jorgenson (1999a & 1999b) has published results for both willow and Miscanthus
that show that, even if organic manures are applied at rates well above standards for
Denmark, then in established plantations nitrate leaching is not enhanced. What we
do not know is what will happen when the plantations are returned to agriculture or
whether in this instance how much of the applied nitrogen was lost by volatilisation
as ammonia. Hall et al. (1996) suggested that the build-up of soil organic matter,
and thus soil nitrogen, would be beneficial in terms of soil structure and moisture
holding capacity when the land is returned to cropping. They state that the nitrogen
will be only slowly mineralised. A comparison may be made with the situation
where old grassland is ploughed up, though organic matter and soil nitrogen levels
may not be as high unless large quantities of organic material has been ‘disposed’
of on the site. Using the changes in organic matter contents described by Johnson
and Prince (1991), Whitmore et al. (1992) calculated that very large losses of
nitrate occur in the first five years or so after ploughing. If soil nitrogen levels
increase during the growth of biomass crops there is likely to be a large loss of
nitrogen by leaching following the return of the land to normal cropping. There
may also be an increased loss following any harvesting operation as, unless soil
conditions are perfect, the trafficking will act like a cultivation - and cultivation has
been shown to increase nitrogen mineralisation. There will also be a large loss of
carbon, as carbon dioxide, from the soil.
2.1.3 Use of organic manures or wastes
Organic manures have two forms of benefit when added to soil. Firstly, they add
nutrients and secondly, they act as soil conditioners and help improve the moisture
holding capacity of the soil. Nutrients in organic manures from human, industrial
and farmyard animal origin contain nutrients in ‘available’ and ‘slowly available’
forms. Standard tables for many of these products are presented in ‘Fertiliser
Recommendations for Agricultural and Horticultural Crops’ (MAFF, 2000).
If manures are applied prior to SRC planting, the value of phosphate and potash in
the manures can be assessed using total contents. This is true unless soil values are
low, when the available nutrient levels should be used, as is standard practice for
agricultural crops. Available nitrogen contents of organic manures can be assessed
using the ammonium nitrogen contents (and uric acid contents for avian manures).
The amounts which will be available for plant uptake can be predicted by using the
decision support system MANNER (Chambers et al., 1999).
As indicated earlier in this report, losses of nitrogen by leaching in the winter
following planting of SRC are elevated. Also, no nitrogen is required in the
establishment year. Logically, it follows that if organic manures are to be applied
prior to planting, they should contain the minimal quantity of ‘available’ nitrogen.
Analysis would show this as ammoniacal nitrogen (poultry manures should be
analysed for uric acid N as well). Therefore, it is concluded that poultry manures,
animal slurries and liquid digested sewage sludge, which contain a high proportion
of available nitrogen, should not be applied prior to planting or in the establishment
year because it will increase the risk of N loss by leaching.
Van den Berg (taken from a translation with no details of the journal) reported that
annual applications of pig slurry after planting and then annually, in an uncontrolled
11
BIOENERGY CROPS AND BIOREMEDIATION
experiment in the Netherlands, increased the growth of poplars above that from
applications of inorganic fertiliser over a five-year period. Spring and summer
applications were marginally better than autumn applications, indicating that at
least some of the nutrients from the pig slurry were lost, possibly by leaching. In a
second set of experiments using calf slurry, there was also an indication from the
growth parameters measured, that slurry applications in early summer, improved
growth more than mineral fertiliser. This occurred on all sites except one, where
rooting was shallow and may have been damaged by traffic during spreading. Some
of the improvement in growth may have been as a result of the extra water from the
slurry. Water is probably the greatest limitation to growth of biomass crops
(Samson, 1993).
In the USA, Sidle & Kardos (1979) reported elevated levels of nitrate loss
following very high applications of nitrogen in bio-solids (1400 and 3000 kg ha-1).
Aschmann et al., (1992) also reported increased losses in the year following
application of waste water sludge, with greatly reduced losses in the second year,
except where high levels of nitrogen had been applied.
Prior to planting and during the preparation of a site for SRC, an application of a
bulky organic manure, such as well-rotted farmyard manure, sludge cake, or other
material with a low available nitrogen content, can be beneficial. Similar yields of
biomass to those achieved with artificial fertiliser have been recorded in Sweden,
where biosolids have been applied (Ledin et al., 1998). Post-planting applications
of materials with higher available nitrogen contents, particularly in the second year
of growth, would be possible, but care should be taken not to exceed the N
requirements of the crop. There will also be an increased risk of losses of nitrogen
as ammonia (see Section 2.2).
These views are in general agreement with those of Riddle-Black (1998) for
biosolids, though she does not distinguish between materials with high available
nitrogen contents used pre-planting and those with low levels, which is important
for environmental reasons, as discussed earlier. Bulky organic materials tend to
improve soil structure and moisture holding capacity, particularly of light soils, and
this can be advantageous. The nitrogen supplied by these materials will be released
slowly by mineralisation and this has advantages, both from environmental and
nutritional points of view. Applications have in the past, been shown to improve
the quality of unrooted rose cuttings (Johnson, 1977). It would be expected that a
similar effect would be seen with SRC, a view also held by Dawson (1999).
If organic manures are applied, their nutrient content should be taken into account
when deciding whether any manufactured fertiliser is to be applied. Standard
nutrient contents for animal manures are contained in standard texts (MAFF, 1994).
The Code of Good Agricultural Practice for the Protection of Water (MAFF, 1998)
specifies that a maximum of 250 kg ha-1 of organic nitrogen can be applied in any
twelve month period. This figure is used throughout this review, for illustrative
purposes. It should be noted that this figure is for total organic nitrogen and that
available nitrogen levels will depend on the type of organic manure/waste being
applied. For low available nitrogen materials, in non-sensitive catchments, 500 kg
N ha-1 is allowed every two years. With the expansion of nitrate vulnerable zones
(NVZs) in 2002 at least 55% of England will fall within sensitive catchments.
12
BIOENERGY CROPS AND BIOREMEDIATION
Phosphate loss to the environment
Phosphate losses are most likely to occur when water erosion occurs. Water-driven
erosion occurs when there is a lack of vegetation preventing the impact of rainfall
on the soil surface leading to surface flow. Therefore, the danger period with
biomass crops is during the establishment period. Malik et al. (2000) have shown
reduced erosion by growing cover vegetation during this period, but all other
vegetation will compete with the biomass crop for water, which is the major
restraint to yield. Thornton et al. (1998) found variable results when comparing
coppice with conventional agricultural crops, with higher losses than from cotton
but less than from maize. Tolbert & Wright (1998) indicated no differences in
erosive soil loss between coppice and conventional crops in the establishment year,
but give no indication as to what happens once the biomass crop is fully
established. A reduction in erosion following establishment might, however, be
expected as raindrop impact would be reduced by interception by leaves and
branches, particularly so in Miscanthus, and soil surface stability would increase as
organic matter contents increase.
If biomass crops are to be used for
bioremediation, growers will need to decide whether or not to use competing cover
crops to reduce early erosion risks, considering factors like rainfall, soil type and
topography.
Many organic waste materials contain large amounts of phosphate. Typical figures
for total phosphate and other nutrients in animal manures and sewage sludge are
given in Tables 2.6 and 2.7 (MAFF, 2000).
Table 2.6. Typical total nutrient content of livestock manures (fresh weight
basis)
Manure Type
Solid manures
Cattle farmyard manure
Pig farmyard manure
Sheep farmyard manure
Duck manure
Layer manure
Poultry litter
Slurries/liquids
Dairy
Beef
Pig
Dirty water
Separated cattle slurries
(liquid portion)
Strainer box
Weeping wall
Mechanical separator
Dry
matter
(%)
25
25
25
25
30
60
6.0
6.0
4.0
< 1.0
Nitrogen
(N)
Phosphate
(P2O5)
6.0
7.0
6.0
6.5
16
30
kg t-1
3.5
7.0
2.0
5.5
13
25
3.0
2.3
4.0
0.3
kg m-3
1.2
1.2
2.0
trace
Phosphate
applied at
250 kg ha-1 N
146
250
83
211
203
208
100
130
125
kg m-3
1.5
3.0
4.0
1.5
2.0
3.0
ND = No data
13
0.3
0.5
1.2
50
62
85
BIOENERGY CROPS AND BIOREMEDIATION
Table 2.7. Typical total nutrient content and available phosphorus of
sludges (kg t-1 or kg m-3 fresh weight)
Digested liquid (kg m-3)
Digested cake (kg t-1)
Thermally dried (kg t-1)
Lime stabilised (kg t-1)
Dry
matter
Total
nitrogen
Total
phosphate
(%)
(N)
(P2O5)
4
25
95
40
2.0
7.5
35
6.0
1.5
9.0
45
8.0
Phosphate
applied at 250
kg ha-1 N
187
299
321
333
Where applications of these materials are made at the maximum nitrogen rate (250
kg ha-1) recommended in the Code of Good Agricultural Practice for the Protection
of Water (MAFF, 1998), large quantities of phosphate will also be added. This is
particularly the case with sewage sludges. Many other industrial organic wastes
may also have large contents of phosphate. The Water Code suggests that, because
of the increased risk of phosphate leaching, applications of phosphate should not
exceed crop removal once soil indices reach 4/5. If regular applications of organic
wastes are made, phosphate applications will certainly exceed eventual removal in
the biomass crop and result in the elevation of soil available phosphate levels.
Alriksson (1998) suggests yields of Salix spp. can range from 16-54 t DM ha-1 over
a five-year period. Using the highest phosphate content given by Frison (1992)
(Table 2.3), this gives a phosphate off-take ranging from 50 to 167 kg ha-1 over 5
years, far less than would be applied by one application of many organic manures.
This confirms that there is the potential for rapid increases in soil phosphate levels,
and thus potential for phosphate loss by leaching, if regular applications are made.
SORGHAL (1997) indicates that experience in Belgium with Miscanthus is that
phosphate removal in a 20t DM ha-1 crop is less than 30 kg ha-1 per year.
Therefore, if UK codes of practice are to be followed, it can be concluded that
biomass crops cannot be used continually for the disposal of high phosphate content
organic wastes.
2.1.4 Landfill leachate and other urban aqueous wastes
Dilute materials have the advantage of supplying large quantities of water, which
can commonly be the limiting factor to the growth of a biomass crop (Samson,
1993), particularly so when the crop is being grown on poor soils, such as those
found on restored landfill sites. Application can be made using overhead irrigators,
but with some materials this may cause leaf scorch, so trickle irrigation may be the
preferred method. Early results from an experiment where landfill leachate is being
applied to a SRC crop using trickle irrigation are extremely promising (Farrow,
pers. comm.).
14
BIOENERGY CROPS AND BIOREMEDIATION
Apart from the possibility of toxic chemicals arising from material disposed of in
the landfill, the expected main potentially polluting compounds are ammonia,
which is toxic to fish, and salt (NaCl) which is phytotoxic at only moderately high
levels (recent work in Sweden is suggesting that it is elevated sodium levels and not
chloride which are reducing willow growth). These would alter the ecosystem of
fresh water streams, if allowed to contaminate them. Land-based disposal of
leachate is common in practice.
In greenhouse experiments Wong & Leung (1989), in Hong Kong, found that yields
of Acacia confusa were reduced when irrigated with landfill leachate; even when
the leachate was diluted with fresh water. However, Cureton et al. (1991) in
Canada found they could increase the growth of Salix babylonica and hybrid
Populus nigra x maximowiczii by irrigating with landfill leachate in a lysimeter
experiment. The leachate used had an electrical conductivity of 10,000 ms cm-1.
However, chlorosis of leaves was observed and leaf senescence occurred 5-6 weeks
earlier than normal. In what appears to be follow-up work, Shrive et al., 1994
reported on field experiments with the same two species on a clay soil in Ontario.
Further investigations by the same group from University of Guelph concentrated
on the effect of leachate sprays on the foliage of established trees. It was found that
leachate spraying caused a decline in transpiration, but after four years no
significant growth effects had been observed. The abstract of this paper concludes
that “Treatment and disposal of MSW leachates in tree plantation may offer a low
technology, low cost option to municipalities.” Menser et al. 1979 reported that a
number of tree species were leachate-tolerant, but in 1983 there was a significant
mortality rate in all species tested. Hasselgren (1998) has reported enhanced
growth of willow by using landfill leachate in Sweden, that he attributed to the
nutrients contained in the leachate.
A general conclusion (DoE, 1997) is that waste waters with conductivities greater
than 2000-4000 ms cm-1 should not be used for irrigation of trees. Many undiluted
leachates will have higher levels. Many leachates contain high levels of nitrogen,
and willows have been shown to be effective scavengers for this nutrient (Edwards
et al., 1998).
It has also been shown (Wong et al., 1990) that allowing landfill leachate to
percolate through soil removes many of the potentially harmful substances within
the leachate. Thus, any water draining from a site where leachate has been irrigated
onto the land is likely to be potentially less harmful to the environment.
In Sweden, Johansson & Elowson (1997) have shown that drainage water
containing high concentrations of nitrate can be ‘cleaned’ by irrigating to excess on
willow plantations.
15
BIOENERGY CROPS AND BIOREMEDIATION
2.1.5 References
Alriksson, B (1987). Influence of site factors on Salix growth with emphasis on
nitrogen response under different conditions. BioBase European Energy Crops
Internetwork.
Aschmann, S G; McIntosh, M S; Angle, J S & Hill, R L (1992). Nitrogen
movement under a hardwood forest amended with liquid waste water sludge.
Agriculture, Ecosystems and Environment 38. 249-263.
Boerjesson, P (1999). Environmental effects of energy crop cultivation in Sweden.
1: identification and quantification. Biomass and Bioenergy 16. 137-154.
Chambers, B J; Lord, E I; Nicholson, F A & Smith, K A (1999). Predicting
nitrogen availability and losses following applications of manures to arable
land: MANNER. Soil Use and Management 15(3). 137-143.
Chambers, B J & Mitchell, R (1995). Crop Nutrition and Water Relations. In:
Arable Energy Coppice (Eds C Britt, M Heath & M Buckland). ADAS,
Oxford. 170 pp.
Christian, D G & Riche, A B (1998). Nitrate leaching losses under Miscanthus
grass planted on a silty clay loam soil. Soil Use and Management 14(3). 131135.
Christopherson, N (1996). Tending of short rotation forests – USA. In:
Handbook on How to Grow Short Rotation Forests (Eds S Ledin & E
Willebrand). Swedish University of Agricultural Sciences/ International
Energy Agency (IEA), Uppsala, Sweden. 3.9.1-3.9.2.
Cureton, P M; Groenvelt, P H & McBride, R A (1991). Landfill leachate
recirculation: Effects on vegetation vigor and clay surface cover infiltration.
Journal of Environmental Quality 20. 17-24.
DoE (1997). The Potential for Woodland Establishment on Landfill Sites.
Edwards, R R; Greaves, M P & Jackson, M B (1998). The potential for the use
of willows as components of practical buffer zones. Proceedings of Long
Ashton Willow Research Open Day October 1998.
Ford-Robertson, J B; Walters, M P & Mitchell, C P (1991). Production and
economics of wood fuel crops for energy forestry. Proceedings Wood - Fuel
for Thought. ETSU, Harwell, Oxon.. 265-282.
Frison, G (1992). Choice of site and establishing short rotation forestry. In:
Handbook on How to Grow Short Rotation Forests (Eds S Ledin & A
Alriksson). Swedish University of Agricultural Sciences/International Energy
Agency (IEA), Uppsala, Sweden.
Hall, R L; Allen, S J; Rosier, P T W; Smith, D M; Hodnett, M G; Roberts, J
M; Hopkins, R; Davies, H N; Kinniburgh, D G & Goody, D C (1996).
Hydrological effects of short rotation energy coppice.
ETSU
B/W5/00275/REP.
Hasselgren, K (1998). Use of municipal waste products in energy forestry:
highlights from 15 years of experience. Biomass and Bioenergy 15(1). 71-74.
16
BIOENERGY CROPS AND BIOREMEDIATION
Johansson, U & Elowson, S (1997). Cleaning of drainage water from agricultural
land using a willow plantation. Proceedings of NJF Seminar No 270,
Alternative Use of Agricultural Land; Foulum, Denmark.
Johnson, P A & Prince, J M (1991). Changes in organic matter in fen silt soils.
In: Advances in Soil Organic Matter Research (Ed. W S Wilson), Royal
Society of Chemistry, Cambridge, 1991.
Johnson, P A (1995). Utilisation of pig slurry on poplars grown for wood: Effects
of husbandry on nitrogen losses by leaching. Proceedings of the ADAS
Terrington poplar seminar. July 1995.
Johnson, P A (2000). Residual effects of nitrogen applications and weed control
techniques on nitrate leaching in a poplar plantation. Report to the Fourth
PAMUCEAF Management Committee Meeting, Torun, Poland.
Jorgensen, U (1999a). Nitrate leaching from Miscanthus and willow after
application of municipal sludge.
BioBase European Energy Crops
Internetwork
Jorgensen, U (1999b). Nitrate leaching from Miscanthus and willow after
application of animal manures. BioBase European Energy Crops Internetwork
Ledin, S & Willebrand, E (1996) (Eds). Handbook on How to Grow Short
Rotation Forests. Swedish University of Agricultural Sciences/International
Energy Agency (IEA), Uppsala, Sweden.
Ledin, S & Alriksson, A (1996). Choice of site and establishing short rotation
forests -Sweden. In: Handbook on How to Grow Short Rotation Forests (Eds S
Ledin & E Willebrand). Swedish University of Agricultural Sciences/
International Energy Agency (IEA), Uppsala, Sweden. 2.6.1-2.6.12.
MAFF (1998). Code of Good Agricultural Practice for the Protection of Water
(The Water Code). Ministry of Agriculture, Fisheries and Food, London.
MacLoughlin, M S; Pope, P E & Hansen, E A (1985). Nitrogen fertilisation and
ground cover in a hybrid poplar plantation: effects on nitrate leaching. Journal
of Environmental Quality 14 (2). 241-245.
Malik, R K; Green, T H; Brown, G F & Mays, D (2000). Use of cover crops in
short rotation hardwood plantations to control erosion. Biomass and Bioenergy
18. 479-487.
Menser, H A; Winant, W M & Bennett, O (1979). Spray irrigation - a land
disposal practice for decontaminating leachate from sanitary landfills.
Mortensen, J; Nielsen, K H & Jorgensen, U (1998). Nitrate leaching during
establishment of willow (Salix viminalis) on two soil types and two fertilization
levels. Biomass and Bioenergy 15 (6). 457-466.
Riddle-Black, D (1998). Development of a water industry manual for biosolids use
in short rotation coppice. Biomass and Bioenergy 15 (1). 101-107.
Samson, R; Girouard, P; Omielan, J & Henning, J (1993). Integrated
production of warm season grasses and agroforestry for low cost biomass
production. In: Proceedings First Biomass Conference of the Americas:
Energy, environment, agriculture and industry.
Burlington, Vermont.
National Renewable Energy Laboratory, Golden, Connecticut.
17
BIOENERGY CROPS AND BIOREMEDIATION
Shepherd, M A; Stockdale, E A; Powlson, D S & Jarvis, S C (1996). The
influence of organic nitrogen mineralization on the management of agricultural
systems in the UK. Soil Use and Management 12 (2). 76-85.
Shrive, S C; McBride, R A & Gordon, A M (1994). Photosynthetic and growth
responses of two broad leaf tree species to irrigation with municipal landfill
leachate. Journal of Environmental Quality 23 534-542.
Sidle, R C & Kardos, L T (1979). Nitrate leaching in a sludge treated forest soil.
Soil Science Society of America Journal 43. 278-82.
SORGHAL (1997). Agronomic aspects of the Miscanthus crop in Belgium.
BioBase European Energy Crops Internetwork.
Thornton, F C; Joslin, J D; Bock, B R; Houston, A; Green, T H; Schoenholtz,
S; Pettry, D & Tyler, D D (1998). Environmental effects of growing woody
crops on agricultural land: First year effects on erosion and water quality.
Biomass and Bioenergy 15 (1). 57-69.
Tolbert, V R & Wright, L L (1998). Environmental enhancement of US biomass
crop technologies: research results to date. Biomass and Bioenergy 15 (1). 93100.
van Veen, J A; Bretelerm, H; Olie, J J & Frissel, M J (1991). Nitrogen and
energy balance of a short rotation poplar forest system. Netherlands Journal of
Agricultural Science 29. 163-172.
Whitmore, A P; Bradbury, N J & Johnson, P A (1992). Potential contribution of
ploughed grassland to nitrate leaching.
Agriculture, Ecosystems and
Environment 39. 221-233.
Wong, M H & Leung, C K (1989). Landfill leachate as irrigation water for trees
and vegetable crops. Waste Management and Research 7 (4). 311-324.
Wong, M H; Leung, C K & Lan, C Y (1990). Decontamination of landfill
leachate by soils with different textures. Biomed Environment Science 3 (4).
429-442.
18
BIOENERGY CROPS AND BIOREMEDIATION
2.2
GASEOUS LOSSES
JOHN KING
2.2.1 Gaseous losses of nutrients
Of the six major plant nutrient elements (C, H, O, N, P and S) all except phosphorus
(P) have a significant gaseous flux element in their normal cycling through the soilplant-air system. Indeed, for C, H and O this is the main method of exchange, in
the forms of CO2, H2O and O2, during photosynthesis and respiration. No
additional burden on these processes is envisaged due to the addition of organic
waste materials to energy crops, with a proviso on water relations, which is
considered elsewhere in this review. This leaves the elements of nitrogen and
sulphur as possibly having their biogeochemical cycles altered by the addition of
organic wastes, such that increased gaseous flux may occur following their
application to energy crops.
Nitrogen
Nitrogen resources may be lost from the soil/organic waste surface in several forms;
NH3, N2O, NOx and N2, depending on edaphic conditions and the nature of the
organic material applied. The amounts lost as N2O , N2 and NOx are minor and
insignificant in nutrient terms, and the N2O flux will be discussed in the following
section. Of major concern is the potential loss of nitrogen nutrition to the crop
from the efflux of ammonia gas, which is a characteristic of manure applications in
agriculture.
From the earliest proposals that the valuable nutrient potential of manures should be
exploited fully, the problem of post-application losses has also been recognised
(Smith & Chambers, 1993). Leaching losses of nitrogen were shown to be
controlled more by soil conditions and time of application, whereas gaseous losses
occurred within hours and were more a feature of the manure type and its mode of
application. Ammonia volatilisation from manures is both a local pollution,
problem of objectionable odours, and a contribution to the diffuse pollution of
atmospheric nitrogen deposition. It is also responsible for a major loss of nutrients
to the crop, being anything from 7 to 84% of slurry ammonium-N applied (Frost et
al., 1990). Of this loss 40-50% occurs within 6 hrs, 70% within 24 hours, and over
90% within 5 days of application (Smith & Chambers, 1993). Besides manure
factors, soil moisture and environmental conditions (windspeed, temperature and
rainfall) also play a part in determining overall losses (Jarvis & Pain, 1990). The
most effective abatement practice on this loss for slurries, is to inject the slurry into
the soil which reduced losses to 2% of ammonium-N compared with surface
applied losses of 74 & 48% (autumn and spring respectively) (Thompson et al.,
1987). This is now common practice for applications to grassland, but is not an
option open to applications on short rotation coppice sites, or even Miscanthus sites,
where woody and perennial root systems are present.
Ammonia volatilises from manures, and in particular slurries, because free
ammonia (dissolved) is in equilibrium with ammonium ions, and outgasses when
the liquid infiltrates into the soil. The equilibrium is such that NH3 is favoured at
19
BIOENERGY CROPS AND BIOREMEDIATION
the high pHs found in slurry (pH 7-8) (Harrison & Webb, 2001), and volatilisation
will also be aided by higher temperatures and windspeeds (Stevens & Laughlin,
1997). Paradoxically, solid manure does not necessarily emit much lower ammonia
concentrations than slurry, especially if the manure is rich in urine. Menzi et al.
(1997) measured 60% of the total ammonium-N lost as NH3 from solid manures
(compared to 52% for slurry), which constituted 10% of total applied N compared
with the 25% for slurry. Chambers et al. (1997) also found a mean emission factor
of 65% of applied ammonium-N for FYM (35% for poultry manure). It is the lower
proportion of total manure nitrogen which is in the ammonium form in solid
manures, which leads to similar, even slightly lower, NH3 emission rates than
slurries, even though solid manures tend to contain more total nitrogen (Chambers
et al., 1999a).
The fact that emissions are largely governed by such a well defined and easily
measured fraction of manures, over only a brief period, has allowed ammonia
volatilisation to be modeled fairly successfully to determine the nutrient values of
manures for following crops. The ADAS model ‘MANNER’ has been used here to
obtain typical losses of ammonia-N for various manures applied and also that
potentially available to a crop (Chambers et al., 1999a) (Table 2.8). The table
assumes a recommended maximum application rate of 250 kg ha-1 N for each
manure type (as for Tables 2.10-2.17), on the 1st April at the end of soil drainage,
which means that no leaching loss was accounted for. The manures were not
incorporated at all and the model made no adjustment for soil type (only in the
leaching component).
Table 2.8. The calculated ammonia emission and plant available N after
surface broadcast applications of various manures to give 250 kg ha-1 N as
soon as drainage has ceased in the spring after harvest, according to
‘MANNER’.
No
Cattle
addition FYM
Application rate t ha-1 DM
Application rate t ha-1 C
Application rate kg ha-1 N
Pig
FYM
Cattle
slurry
Pig
slurry
Sewage
sludge
0.00
0.00
0.00
10.50
4.73
250.00
9.00
4.05
250.00
4.90
2.21
250.00
2.40
1.08
250.00
8.30
3.74
250.00
Potential plant N kg ha-1
0
81
83
138
160
58
Volatilised ammonia N kg ha-1
Available plant N kg ha-1
0
0
40
41
41
42
46
92
39
121
24
34
As can be seen from Table 2.8 the volatilised ammonia loss of potentially available
nitrogen to the crop is considerable - between 24 and 46 kg ha-1 for all cases; which
constitutes between 32 and 98% of that which is eventually available to the current
crop. The figures in this table are also derived from rather benign conditions. If
manure is applied earlier in the season, at the end of February or early March, then
losses will also be exacerbated by leaching.
20
BIOENERGY CROPS AND BIOREMEDIATION
2.2.2 Emissions of radiatively active gases from soils and organic
waste materials
The ‘greenhouse effect’ - whereby increased concentrations in the troposphere of
certain gases from anthropogenic sources cause the retention of infra-red radiation
and thereby global warming of the atmosphere (Dickinson & Cicerone, 1986) - is
now a well documented and accepted phenomenon (IPCC, 1997). Concern over
this is high enough for international agreements to have been sought limiting
emissions of the major gas concerned, CO2 (The ‘Kyoto protocols’). This concern
is a major reason for the perceived viability of bioenergy crops as an increasing
component of the UK’s energy generation programme - as they are believed to be
‘carbon neutral’ (DTI, 1999). This means that although their combustion in power
stations will generate CO2, this will merely re-cycle that which they previously
fixed from the atmosphere during photosynthesis, and will once again be fixed by
succeeding crops.
When considered on a ‘per unit area of land’ basis, this expected neutral balance of
carbon input and output will be radically altered by the addition of organic wastes
as fertilisers to the crops, during early stages in the rotation. The use of such
materials also complicates the picture in that they are known to emit other
‘greenhouse gases’ (depending upon application and edaphic conditions), which are
more effective than CO2 at trapping infra-red radiation. The gases in question are
chiefly methane (CH4) and nitrous oxide (N2O), which have approximately 32 and
150 times the potential to absorb infra-red radiation respectively, as CO2
(Bouwman, 1990). Houghton et al. (1996) however, put the effect due to N2O as
high as 280 times that for CO2, and more recently Smith et al. (2001) quoted the
same value for N2O and a CH4 forcing value of 56 CO2 equivalents.
All of the gases are associated with organic wastes, though the actual flux emitted
will vary according to the nature of the material, edaphic conditions and the way in
which it is applied to the soil. Both N2O and CH4 are produced by soil processes
that take place under anaerobic conditions, so methods of application that minimise
the soil air-space or gaseous exchange will exacerbate their emission. Similarly, the
application of very large amounts can lead to anaerobic zones being created.
The factors that govern the emission of the two main gases and their possible
emission rates will be considered here in separate sections for each gas. Their
likely emissions following the application of the major organic waste materials will
be considered in turn, and a final section will discuss the likely perturbation to the
CO2 flux cycle during the crop rotation, and overall carbon dioxide equivalent
balance per unit area of land.
Evolution of nitrous oxide (N2O)
The production of N2O in soils is a result of natural soil processes and occurs in all
soil types at a low level. It is both a minor by-product of the ubiquitous process of
nitrification, and also a major product of the process of denitrification which only
occurs under anaerobic conditions (Bouwman, 1990).
21
BIOENERGY CROPS AND BIOREMEDIATION
During nitrification NO2 is produced in the first reaction step mediated by
Nitrosomonas species, according to Equation 1 - before further oxidation by
Nitrobacter species to the nitrate ion.
Equation 1
NH4+ + 3/2 O2  NO2- + 2H + H2O + E
Some of the NO2-, however, is reduced to N2O by nitrifying bacteria, when oxygen
is partially limiting or soil acidity is high (Firestone & Davidson, 1989). Therefore
some flux of N2O from soil under energy crops should be expected following the
ammonification of organic N in plant litter. The magnitude, however, is unknown as emission rates under willow, poplar or Miscanthus plantations have not been
measured - but would probably be of the order of 5-25 g N ha-1 d-1, similar to that
found under beech (Fagus sylvatica) and alder (Alnus) forests by Mogge et al.
(1996) (although denitrification will also contribute to this). Any activity that
increases the ammonium concentration in the soil will tend to increase N2O flux
during its subsequent nitrification. Therefore, the application of urea and
ammonium salt fertiliser will exacerbate this pathway for N2O loss if the edaphic
conditions are favourable. The application of nitrogenous organic waste materials
to bioenergy plantations may also constitute a net increase in N2O loss by
nitrification, but this will be spread over a long period of time, as the material
decomposes, and may not be significantly above background rates of emission.
Several factors linked to the application of inorganic waste material to soil favour
the more productive formative pathway for N2O in the soil, that of denitrification.
In this process anaerobic conditions lead to NO3- ions in soil being used as an
electron donor for respiration of certain facultative anaerobic micro-organisms. The
general reduction pathway for this process is that in Equation. 2, in which N2O is an
intermediate result, before the full result of the pathway to molecular nitrogen
(Bouwman, 1990). As such it may be emitted from the soil surface, alongside
nitrogen gas.
Equation 2
NO3-  NO2-  NO  N2O  N2 + H2O
The microbial requirements for this process are the lack of molecular oxygen,
available NO3- ions as a substrate, and an accessible carbon source for energy
(Dendooven et al., 1997). Naturally denitrifying organisms must also be present,
and other controlling factors are the temperature and pH of the soil medium. The
full denitrification process is not a problem for climate change issues, as N2 is
radiatively neutral; but factors which increase the ratio of N2O:N2 emission, such as
lowered pH, increased access to simple sugars as a C source and periods of
anaerobicity, all act to repress NO2- reduction and so lead to a greater efflux of N2O
relative to N2 (Dendooven et al., 1997). The emission rate of N2O was reduced by
a decline in pH from 6 to 4, because of an increased lag time before NO2- reduction
progressed (Ellis et al., 1997). However, below pH 4 denitrification effectively
ceases and N2O production in acid forest soils at these low pHs is due mainly to
nitrification processes. Although SRC species will tolerate a wide range of soil pH,
their yield is reduced on acid soils, and liming to a range of 5.5-8.0 is advised for
planting (Britt et al., 1995). Denitrification is therefore unlikely to be inhibited by
pH on SRC sites.
22
BIOENERGY CROPS AND BIOREMEDIATION
The addition of an organic waste to soil will not only increase the nitrate nitrogen
available for denitrification (after mineralisation and ammonification from the
organic material), but also the available carbon substrate required. Furthermore, the
presence of large quantities of water, such as slurry wastes provide, and the active
consumption of oxygen by mineralisation processes, will increase the likelihood
and extent of anaerobic sites for denitrification within the soil (Dobbie et al., 1999).
This last effect will be exacerbated by the soil compacting activity of manure and/or
slurry transporting and spreading machinery (Smith & Chambers, 1993), and be
most damaging on already compacted soils such as derelict land or clay soils
susceptible to the formation of plough pans. The very act of harvesting the crop in
the winter period will lead to compaction on clay sites (Britt et al., 1995), and
therefore the conditions during the next rotation will be such that losses of N2O
after organic waste application are likely to be high. Just how high is again open to
speculation though, as no studies quantifying this loss have been found in the
literature, and the following sections extrapolate emission rates from other land use
situations.
Livestock slurries
The application of cattle and pig slurry wastes to AEC land can be seen as a
sensible re-cycling of waste material from intensive livestock units with
considerable nutrient value to the crop. Slurries can contain of the order of 0.330.44% N, 0.15-0.22% phosphorus and 0.41-0.78% potash on a fresh weight basis
(Smith & Chambers, 1993). Approximately half of the nitrogen is in the
ammonium form, and considerable quantities can be lost as ammonia gas.
Nitrogen lost as N2O from clayey sand and sandy loam soils in Denmark was
increased twofold by the application of pig slurry compared with mineral nitrogen
fertiliser (Maag et al., 1996). The increase amounted to 1-3 kg ha-1 a-1 N for
additions of 25-36 kg ha-1 N as organic N, when 60-90 kg ha-1 N was also applied in
mineral form. In earlier studies in Denmark by Christensen (1983), the proportion
of applied N lost as N2O, rose from 8.2% when applied as NH4NO3 (200 kg ha-1 N),
to 37% when applied as cattle slurry (492 kg ha-1 N) (in Fowler et al., 1997). Over
a year this amounted to 182 kg ha-1 N lost as N2O.
Emissions of N2O from either mineral or slurry additions to soil are highest
immediately after application. Peak values of up to 1.6 kg ha-1 d-1 N were found on
a Scottish clay loam by McTaggart et al. (1997), though this fell within 10 days to
background levels. In their studies, McTaggart et al. (1997) showed that, when
compared with slurry, mineral fertiliser nitrogen was actually the greater source of
N2O and that soil compaction was the biggest factor in increasing emissions.
However, although slurry in addition to NH4NO3 meant no increase in N2O flux on
two occasions, an additional 1.5-2.6 kg ha-1 N flux was apparent in the 60 days after
application on the other two occasions (spring and summer), when 75 kg ha -1 N was
applied as slurry. In general, however, the N2O loss after the application of slurry
constitutes a higher proportion of applied ammonium-N (0.1-4.0%) than that from
mineral-N fertiliser (0.1-0.9%) (Eichner, 1990; Harrison & Webb, 2001).
The timing of slurry applications has proved a most influential factor in controlling
the scale of N2O emissions. Generally emissions (N2O + N2) in the autumn or
23
BIOENERGY CROPS AND BIOREMEDIATION
winter ranged from 23-29% of applied ammonium-N compared with 4-5.5% for
spring or summer applications (Stevens & Laughlin, 1997). In the autumn and
winter, nitrate-N formed by the nitrification of ammonium-N in slurry is not
removed by crop uptake and is thereby open to denitrification. In the spring and
summer, crop uptake removes nitrate from potential denitrification, and soil
moisture deficits also mean anaerobic conditions are less likely. However, after
heavy rain or in compacted soil conditions, the higher soil temperatures of summer
can lead to high emission rates (McTaggart et al., 1997).
Emission rates of N2O (and N2) from slurry have been shown to demonstrate a trend
of flux proportional to application rate (Paul et al., 1993), similar to the relationship
for mineral N application (Eichner, 1990). However, evidence from field trials is
less clear, but supports the presumed trend (Velthof & Onoema, 1993).
Measured emission rates from soil amended with slurry are difficult to summarise,
due to variable application rates and methods, additional treatments and fertiliser
additions, and a range of alternative measurement techniques used. However, Table
2.9 is an attempt to quantify the emissions found relative to the amount of slurry
and ammonium-N applied. The figures vary widely, but for typical agricultural
total nitrogen applications of 25-125 kg ha-1 in one application, or split into two, an
estimated 0.5-15 kg N is lost as N2O in the year after application, broadly in
proportion to the amount applied. These figures are generally in agreement, though
tending to be higher than the 2% “rule of thumb” value Eichner derived for the
proportion of applied manure N lost as N2O. However, in their UK inventory of
N2O emissions from farmed livestock, Chadwick et al. (1999) quoted emission
factors of 0.4% for pig slurry and 0.3% for cattle slurries, which is only a quarter of
those quoted previously.
Solid livestock manures
Fewer studies have been made of emission rates from solid manures. Although
ammonia losses from such manures can be very high, the drier nature of the
material means that denitrification, and therefore high N2O losses, are less of a
problem. Nevertheless, if heavy rain follows manure application then N2O fluxes
can be significant. Coyne et al., 1995 report such a scenario for poultry manure
applied to tilled soil at the high rate of 448 kg ha-1 N and measured fluxes of 1.3-3.2
kg ha-1 d-1 N2O-N loss, whereas Cates & Keeney (1987) measured annual losses of
3.6-5.2 kg ha-1 N2O-N. Both of these values appear high compared to those in
Table 2.9, but are for very high application rates.
Lessard et al. (1996), on the other hand, measured N2O fluxes that amounted to
only 1% of applied dairy manure nitrogen under maize. They found efflux rates of
4.1 x 10-2 to 1.2 x 10-1 kg ha-1 d-1 N2O-N from applications of 170 and 339 kg ha-1
N respectively. These were only significantly above background rates in the first
seven days after application and 67% of total emissions occurred during the first
seven weeks.
Set against the general picture from short-term studies, which measure the highest
fluxes in the initial stages after application, is the study of Chang et al. (1998),
which measured emissions over one year from plots which had received manure
24
BIOENERGY CROPS AND BIOREMEDIATION
over 21 years previously. These barley plots had had cattle feedlot manure applied
and incorporated every year at rates from 60-180 t ha-1 (as well as 0), and the
annual N2O emissions ranged from 11 to 56 kg ha-1 a-1 N2O-N (0.7 kg ha-1 a-1 in the
0 plots). At dry matters of 58% and total N contents of 1.6% dry matter, these
amounts again accounted for approximately 2-3% of the total N applied to the plots.
A warning is given by this study, that prolonged application of manures can raise
emissions, and that predictions based on short-term studies may underestimate
efflux in such situations. Chadwick et al. (1999) quoted a seemingly high figure of
5.9% N loss from solid manures when expressed as a percentage of ammonium-N,
but as a proportion of total N (typically 10% for solid manures, Menzi et al., 1997)
is only 0.6%, which is more in line with their conservative estimates of flux quoted
for slurries.
Sewage sludge
Despite the huge amount of work that has been done on the impact that sewage
sludge has on soil microbial processes when applied to land (Smith, 1991), very
little is reported in the literature about emissions of greenhouse gases after
application. Digested sludge cake has a broadly similar nitrogen content to
farmyard manure (Aitken, 1997), and so can be expected to behave similarly in
stimulating N2O emission from land by providing both nitrogen and carbon
resources to soil nitrification and denitrification processes. In a study where large
amounts of sludge cake (16.7 and 83.5 t ha-1 DM) were applied to a sandy loam soil
in an arable rotation the mean daily N2O efflux was measured over a five month
period from a sown barley crop by Mosier et al. (1982) in Colorado, USA, and
compared to similar amounts of N applied as ammonium nitrate fertiliser. At the
lower rate (corresponding to 71 kg ha-1 N) rates were of the order of 4 g ha-1 d-1
N2O-N above the control, slightly higher than that for 56-112 kg ha-1 N applied as
ammonium nitrate (about 2.5-3.5 g ha-1 d-1). At the higher rate (356 kg ha-1 N),
however, 2.2 g ha-1 d-1 N2O-N was emitted above the background, much in excess
of the 6 g ha-1 d-1 emitted after 224 kg ha-1 of ammonium nitrate N. Over the full
period this amounted to emissions of 0.8 and 1.0% of the applied N evolved as
N2O. As for livestock wastes, emissions were highest in the initial post-application
period and 69% of those over the 155 day measurement period evolved during the
first six weeks. Peaks after this followed fluctuations in soil water content.
The effect of rainfall events in leading to peaks in N2O emission from grassland
after sewage sludge application, was shown in the only reported UK study by Scott
et al. (1998). The full report of this experiment (Scott et al., 2000), details the
emission pattern after rates of 185 t DM ha-1 sludge (1.37% N) had been applied for
three years (2,535 kg ha-1 a-1 N) to grassland, on an imperfectly drained sandy clay
loam soil in southern Scotland. Initial peak diurnal fluxes of 48 to 120 g ha -1 d-1
N2O-N were experienced, which declined to about 10 g ha-1 d-1 above background
after about 30 days. This was punctuated though by similar order peaks, following
rainfall events, to those initially experienced, illustrating the controlling influence
of the interaction between soil permeability and incident rainfall. The cumulative
total efflux over a period of six months from August to March, was 10-20 kg ha-1
N2O-N above that from plots fertilised at a similar rate with ammonium nitrate (480
kg ha-1 a-1 N) and represents 0.4-0.8% of the N applied as sludge.
25
BIOENERGY CROPS AND BIOREMEDIATION
Table 2.9
Nitrous oxide emission rates from livestock slurry manures cited in the literature
Soil type
Additional
features
Slurry type
Sandy loam
Grassland
Dairy
Amount of
total N
applied (kg
ha-1 d-1 N)
492
Sandy loam
Grassland
Cattle
200
Sandy
Grassland
Dairy
1.4-4.2
Sedge Peat
(38% OM)
Clayey sand
Grassland
Cattle
0.7-0.9
Arable
Cattle
36-45
Sandy loam
Clay loam
Arable
Imperfectly
drained
grassland
Free draining
grassland
Arable
unknown
25-54
44-75
Dairy
44-124
Sandy loam
Sandy loam
Sandy loam
over Loamy
sand
Silty clay
loam
Sandy
Amount
applied (t ha-1
DM)
45.1-50.1
Pig
Amount of
NH4+-N
applied (kg
ha-1 d-1 N)
27.7-72.9
9.6
24.3-25.6
11-380
Pig
100-162
Arable
Pig
225
Arable
Cattle & Pig
80-100
Long term
flux rate (g
ha-1 d-1 N)
Estimated
flux over one
year (kg ha-1
a-1 N)
182
0.23-530
0.4-5.3
11-200
18.2
1-5
1-85
1-2
0.4-31
0.56-2.2
0.2-0.8
25-42
9.1-15.3
2.5-12
0.52-1.82
30
10
0.3-1.3
5-25
3
1.1-1.8
0.2-1.6
32-56
66-90
Arable
Short term
N2O flux rate
(g ha-1 d-1 N)
12-180
83-96
52-65
26
Reference
Christensen,
1983
Egginton &
Smith, 1986
Velthof &
Onoema 1993
Jars et al.,
1994
Maag et al.,
1996
“
McTaggart et
al., 1997
Chadwick,
1997
Ferm et al.,
1999
Weslien et al.,
1998
Arcara et al.,
1999
Petersen 1999
BIOENERGY CROPS AND BIOREMEDIATION
The above evidence suggests that over a full year the emission of nitrogen
as N2O is likely to be of the order of 1-2% of that applied, similar to
livestock wastes.
Evolution of methane
Methane (CH4) is also produced in soils under anaerobic conditions, when
the inorganic hydrogen acceptor ions such as the nitrates have been used up
(section 2.1) and the fermentation of organic molecules commences
(Bouwman, 1990). Nitrates repress this formation by both delaying the
onset until they have been reduced, but also by direct toxicity to
methanogenesis. Sulphates will also act similarly. Methane production
after the application of large quantities of organic wastes to soil, will
therefore only be likely under prolonged anaerobic conditions.
Methanogenesis can occur mainly by two processes, either using acetates as
a substrate according to Equation 3, or by the reduction of CO2 as in
Equation 4 (Batjes & Bridges, 1992):
Equation 3
Equation 4
CH3COOH > CH4 + CO2
CO2 + 4H2 > CH4 + 2H2O.
Methane is also widely consumed by methanotrophic organisms in aerobic
soil (Batjes & Bridges, 1992) and by some ammonium oxidising
organisms. Aerobic soil can therefore act as a sink rather than a source of
methane (Schütz et al., 1990), its presence inducing greater growth and
activity of methanotrophic organisms (Bender & Conrad, 1995). For these
reasons methane emissions are usually only associated with natural
wetlands or rice paddy soils where anaerobic conditions are maintained as
an equilibrium, and their effects outweigh that of any aerobic topsoil.
Certain situations such as landfill sites where large quantities of organic
material are buried under anaerobic conditions also become net methane
emitters.
Methane from livestock and sewage wastes
The huge production of animal and human waste also contributes to the
global budget of atmospheric methane, both directly from ruminant animals
in the field, but also the storage and processing of waste materials.
Generally livestock wastes are highly conducive to methane generation,
containing both the organic substrates required and suitable active microorganisms. The actual amounts produced are however highly dependent on
the diet of the producer animals and the way in which the manure is treated.
Grain-fed housed animals have been found to produce more methane than
those kept extensively on low grade forage (Hogan, 1993), and Jarvis et al.
(1995) found an inverse relationship between production and the C:N ratio
of cattle dung. The main controlling factor in the handling of livestock
wastes for methane production is undoubtedly the way in which it is stored,
and liquid based storage systems (slurry, etc.) produce the most. Contact
with oxygen will suppress methane production for the reasons given above,
27
BIOENERGY CROPS AND BIOREMEDIATION
as will extremes of pH, because methanogens require conditions around
neutrality (pH 6.6-8.0; Hogan, 1993).
Hardly any citations of methane emissions from field-spread manures were
found (Sommer et al. (1996) (and none for sewage sludge) though some
workers had measured the efflux from cattle and sheep dung (Lodman et
al., 1993; Williams, 1993; Jarvis et al., 1995). These studies produced
broadly comparable results, that emissions were highest immediately after
deposition and lasted over a period of about 10 days, declining with
increasing aerobicity of the dung. Values ranged from about 300-2000 mg
CH4 m-2 in the field, varying according to type of animal and feeding
regime; but amounting roughly to a mean of 74 mg kg-1 wet weight of
application. By contrast Willison et al. (1996) found that additions of FYM
to soil increased the methanotrophic activity by ammonium fertilisation
from mineralising organic nitrogen! Any emissions from applied manure
or sewage sludge are therefore likely to be small - highly dependent on both
the amount and type of material applied and the soil hydrology immediately
after application, and limited to the immediate post application period (1014 days). Indeed, Sommer et al. (1996) only measured 30 g of CH4-C ha-1
emitted from pig slurry spread on bare soil, and all of that within the first
day after application. Neither Sommer et al. (1996), nor Jarvis et al. (1995)
found that the soil hydrology altered emissions, which suggests that it is
only a temporary event dependant entirely on the conditions of the manure,
supporting the view that some at least is dissolved CH4 volatilising out of
the manure, rather than true methanogenesis (Sommer et al. (1996).
A predictive emission rate is impossible to give and only estimates for
livestock wastes can be attempted. National and international inventories
for greenhouse gas emissions (Hogan, 1993; IPCC, 1997) tend to estimate
emissions across the entire manure-handling chain. FYM drylot storage
and slurry/lagoon systems with higher emission factors produce most
methane during the storage phase. Although it should be noted that the use
of manures entails this necessary step, and thereby CH4 emission, this
would happen regardless of their use on energy crop sites. Consequently, it
should not affect the argument about carbon flux mitigation by energy
cropping being reduced by on-site greenhouse gas emissions. However, to
produce an estimate for this report we have used generic figures for the
amount of volatile solids (VS) in all types of fresh manure (10-12% dry
matter) and a methane conversion factor (MCF) for the daily spread or
pasture manure handling systems (0.1-2%, over 10-300C range), taken from
Hogan (1993) (from the work of Hashimoto, also used by the IPCC, 1997).
Soil carbon balance
The amount of carbon added to the soil in various organic wastes will, of
course, depend upon their composition and the amounts applied. However,
the mineralisation of these organic resources follows a fairly predictable
pattern of the mineralisation of any vegetation derived material, described
roughly by an exponential relationship (Jenkinson, 1990). As a rule of
thumb, 66% of added vegetable material is lost as CO2 within a year of
28
BIOENERGY CROPS AND BIOREMEDIATION
incorporation.
During this year a higher proportion will be lost
immediately after incorporation and a smaller proportion will be cycled
through the soil microbiota and into the soil organic matter (SOM). In
agricultural soils whose organic matter content is at equilibrium, this
amount added replaces that lost from mineralisation of soil organic matter
for annual additions. Models for carbon turnover exist, such as that of
Jenkinson (1990), which can be used to predict the decline in added carbon
(and the new soil organic matter equilibrium) over a crop rotation, but an
approximation of 10% has been used here. This represents the amount
remaining after two years at the initial rate of loss, but it is recognised that
this rate declines exponentially as easily decomposed material is respired
and only recalcitrant matter remains, and 10% is seen as a median value for
recalcitrant material which may remain for several years.
In addition to any increase in equilibrium soil organic matter level due to
organic waste application, there will be a small increment from litter
deposition and turnover. Litter deposition from willows grown on a
Swedish peat bog showed slightly retarded decomposition dynamics
compared to the “rule of thumb” mentioned above, losing about half their
mass over three years and only declining at a very low rate after that
(Slapokas and Granhall, 1991). Given that temperatures will be lower and
the peat surface wetter in their study than most UK mineral soil sites, we
can still assume that up to half the deposited litter carbon will still comprise
a deep litter mulch on the soil surface after three years, at the end of the
harvest rotation. If leaf litter constitutes 30-40% of the above ground
woody biomass, as is the case for alder plantations (Rytter et al., 1989),
then willow crops producing a typical annual yield increment of 12 t ha -1
DM will deposit about 11-14 t ha-1 DM as litter over a three year rotation of which half may constitute a standing litter layer from the first harvest
onwards (about 5-7 t ha-1 DM). This layer, and the turnover of fine roots,
will also contribute an estimated 4-10 t ha-1 DM annual turnover, leading to
a small increment in soil organic matter of the order of 0.5 t ha-1 a-1 carbon
(Börjesson, 1999).
There will also be a net increase in below ground carbon during the first
energy crop rotation, caused by the development of a semi-permanent
woody root system. In most trees the root system can be the sink for a
considerable proportion of the annual carbon increment; for example 11%
for fine roots and 31% for coarse roots were recorded in a grey alder
plantation in Sweden (Elowson & Rytter, 1993), but the coarse root
standing biomass tends to stabilise around 20% over all species (Ericsson et
al., 1996). Using this mean figure, and typical UK willow annual biomass
production of 12 t ha-1 a-1 DM (Bullard, pers. comm.), then 2.4 t ha-1 DM
will be deposited after one year and 9.6 t ha-1 DM at the first harvest of
willow after four years. This latter figure is likely to be the standing mass
throughout the remainder of the 25 years of the full economic rotation of an
energy crop plantation. The energy grass Miscanthus apportions an even
greater component of carbon allocation to its root system in the initial
building phase of growth, putting about 1 t ha-1 DM in the first year and 810 t ha-1 DM after two years, until it develops a standing root biomass of
29
BIOENERGY CROPS AND BIOREMEDIATION
about 15 t ha-1 DM in the third year. The root mass remains at this level for
the remainder of the 25 year plantation, for a crop delivering 15 t ha-1 a-1
DM of shoot material (Bullard, pers. comm.).
The carbon-equivalence balance of a site
Using guide figures from the literature detailed above, the net effects of
typical energy crops and organic waste fertiliser practice, on the
contribution to radiatively active gases in the atmosphere, and carbon
sequestration, can be estimated on a site basis. To do this, one mole of
methane has been taken as being equivalent to 56 moles of CO2 and one
mole of N2O as being equivalent to 280 moles of CO2 (Smith et al., 2001).
Increments of carbon into a site are considered a positive gain to carbon
sequestration, whilst losses of greenhouse gases are considered a loss to
carbon sequestration value, and for all organic waste and vegetation
materials the carbon concentration has been taken as a universal mean of
45% DM (Palm & Rowland, 1997). Livestock wastes and sewage sludge
have been assumed to be applied at the maximum loading allowed for
arable land which gives an equivalent of 250 kg ha-1 N (Chambers et al.,
1999b) using typical nitrogen contents (Anon, 2000). For cattle and pig
FYM and digested sewage sludge cake, this means application rates of
10.5, 9.0 and 8.3 t ha-1 DM respectively, and for cattle and pig slurry, 4.9
and 2.4 t ha-1 DM.
The suggested rotation for willow coppice is that they are cut after one year
(the cuttings left on site), when they can be fertilised by an application of
organic waste material, and thereafter every three years until the full
economic rotation of 25 years has elapsed. Organic waste materials would
be applied after every three-year harvest. The carbon sequestration
potential for such a site is calculated in Table 2.10 for the first year, Table
2.11 for the first harvest period, and in Table 2.12 for the full 25 year
period.
In the first year with no waste application, all plots register a small positive
contribution to carbon sequestration on the site (Table 2.10). In the fourth
year after one application of waste, there has been a net contribution of
carbon to the site by all treatments, but the effect of potential greenhouse
gas emissions has been to reduce the effectiveness of this to a negative
impact, relative to a site with no addition (Table 2.11). Over the full
economic rotation this effect becomes even more marked, reducing the
benefits of carbon sequestration by between 27 and 36%.
A similar scenario for Miscanthus, gives the values in Table 2.13, 2.14 &
2.15 respectively, again only applying organic wastes in every third year,
even though Miscanthus is harvested every year. The amount of carbon
sequestered is less than willow in the early stages (due mainly to no
standing litter being accounted for, and a smaller root system), but the same
effect of greenhouse gas emissions reducing the effectiveness of this
sequestration is also apparent (Tables 2.12 & 2.13). After 25 years the
30
BIOENERGY CROPS AND BIOREMEDIATION
carbon sequestered is slightly greater than willow coppice, and again its
effectiveness is reduced by 24-31% (Table 2.13).
If wastes are applied at every opportunity then the carbon sequestration
potential increases considerably, even after only four years (Table 2.16).
However, the attenuation of this by greenhouse gas emission also increases,
as can be seen from the 30-39% reduction in effectiveness displayed in
Table 2.16. Over the full 25 years, this amounts to large quantities of extra
carbon in the topsoil (Table 2.17), but a 68-75% negation of the benefits of
this with respect to greenhouse gas production. This is because, all the
losses of N2O and CH4 occur in the initial post application period and, in
the annual application scenario of Table 2.17, the annual emission of these
gases overwhelmingly outweighs the minor accrual of soil carbon.
What is not taken into account in the balances given in Tables 2.10-2.17 is
the CO2 emissions from the 90% of applied organic waste that mineralises,
nor the carbon in the harvestable biomass of crop. The crop biomass is
considered to be entirely neutral in that the same amount of carbon taken
out of the atmosphere in its formation is returned on incineration (and
through plant respiration). The CO2 respired during organic waste
respiration is a greenhouse gas emitted from the site, but it was considered
that this would have been returned to the atmosphere whatever the
treatment of the material either by mineralisation or incineration. The
accrual of soil organic matter from its application is a genuine, though
temporary, sequestration of carbon to the site which would not otherwise
occur, and the contribution from N2O and CH4 to greenhouse gas
emissions, are considered to be different to CO2, as their production is more
dependant upon handling and management issues, and technically need not
occur if perfectly aerobic conditions could be maintained.
The estimates in Tables 2.10-2.12, of the impact of various waste additions
on the balance of carbon sequestration and greenhouse gas emissions from
an SRC site, are based upon very rough figures from work in normal
agricultural contexts. As such, they may not necessarily reflect the
situation when large amounts of waste are disposed of to coppice sites on
reclaimed land, where soil compaction by reclamation plant or large-scale
harvesting equipment may create a soil hydrology that exacerbates the
likelihood of greenhouse gas production. Nor do they take any account of
other possible waste applications that involve irrigation (by waste water or
landfill leachate, etc.) on a regular basis, which again will create conditions
conducive to greenhouse gas emissions.
Of particular concern would be the annual application of waste to
Miscanthus, which is likely to produce even worse conditions than
suggested in Tables 2.16 and 2.17 - as the accrual of organic matter from
the waste is probably under-estimated in these tables, and this would act to
retain moisture and make anaerobiosis more frequent and extensive.
With these provisos however, the practice of applying moderate amounts of
organic wastes to energy crops is likely to be both beneficial to the soil
31
BIOENERGY CROPS AND BIOREMEDIATION
sustainability of the site and make a modest contribution of carbon
sequestration to the overall balance of factors which contribute to the
forcing of climate change.
32
BIOENERGY CROPS AND BIOREMEDIATION
Table 2.10. Carbon equivalent fluxes (C t ha-1) on a willow coppice site one year
after planting.
1st year - willow
No
addition
Cattle
FYM
Pig FYM
Cattle
slurry
Pig slurry Sewage
sludge
Application rate t ha-1 DM
Application rate t ha-1 C
Application rate kg ha-1 N
0.00
0.00
0.00
10.50
4.73
250.00
9.00
4.05
250.00
4.90
2.21
250.00
2.40
1.08
250.00
8.30
3.74
250.00
C in litter (6 t ha-1 DM)
C in roots (2.4 t ha-1 DM)
C in SOM (0.5 t ha-1 yr-1)
2.70
1.08
0.50
2.70
1.08
0.50
2.70
1.08
0.50
2.70
1.08
0.50
2.70
1.08
0.50
2.70
1.08
0.50
N20 efflux (t ha-1 CO2-C equivalent)
CH4 efflux (t ha-1 CO2-C equivalent)
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
Total C sequestered
Site balance of C sequestration
4.28
4.28
4.28
4.28
4.28
4.28
4.28
4.28
4.28
4.28
4.28
4.28
Table 2.11. Carbon equivalent fluxes (C t ha-1) on a willow coppice site after one
harvest, four years after planting, with organic wastes applied after one year.
Application rates the same as Table 2.10.
4th year - willow
No
addition
Cattle
FYM
Pig FYM
Cattle
slurry
C in litter (6 t ha-1 DM)
C in roots (9.6 t ha-1 DM)
C in SOM (0.5 t ha-1 yr-1 )
C from waste to SOM (10%)
2.70
4.32
2.00
0.00
2.70
4.32
2.00
0.47
2.70
4.32
2.00
0.41
2.70
4.32
2.00
0.22
2.70
4.32
2.00
0.11
2.70
4.32
2.00
0.37
N20 efflux (t ha-1 CO2-C equivalent)
CH4 efflux (t ha-1 CO2-C equivalent)
0.00
0.00
0.60
0.44
0.60
0.37
0.60
0.20
0.60
0.10
0.60
0.35
Total C sequestered
Site balance of C sequestration
9.02
9.02
9.49
8.46
9.43
8.45
9.24
8.44
9.13
8.43
9.39
8.45
33
Pig slurry Sewage
sludge
BIOENERGY CROPS AND BIOREMEDIATION
Table 2.12. Carbon equivalent fluxes (C t ha-1) on a willow coppice site after the
full economic rotation of 25 years, with organic wastes applied after one year, and
again after every three year harvest. Application rates the same as Table 2.10.
25th year - willow
C in litter (6 t ha-1 DM)
C in roots (2.4 t ha-1 DM)
C in SOM (0.5 t ha-1 yr-1)
C from waste to SOM (10%)
N20 efflux (t ha-1 CO2-C equivalent)
CH4 efflux (t ha-1 CO2-C equivalent)
Total C sequestered
Site balance of C sequestration
No
addition
Cattle
FYM
Pig FYM
Cattle
slurry
Pig slurry Sewage
sludge
2.70
4.32
12.50
0.00
2.70
4.32
12.50
3.78
2.70
4.32
12.50
3.24
2.70
4.32
12.50
1.76
2.70
4.32
12.50
0.86
2.70
4.32
12.50
2.99
0.00
0.00
4.80
3.49
4.80
2.99
4.80
1.63
4.80
0.80
4.80
2.76
19.52
19.52
23.30
15.01
22.76
14.97
21.28
14.85
20.38
14.79
22.51
14.95
Table 2.13. Carbon equivalent fluxes (C t ha-1) on a Miscanthus plantation one
year after planting.
1st year - Miscanthus
No
addition
Cattle
FYM
Pig FYM
Cattle
slurry
Application rate t ha-1 DM
Application rate t ha-1 C
Application rate kg ha-1 N
0.00
0.00
0.00
10.50
4.73
250.00
9.00
4.05
250.00
4.90
2.21
250.00
2.40
1.08
250.00
8.30
3.74
250.00
C in litter (0 t ha-1 DM)
C in roots (0.8 t ha-1 DM)
C in SOM (0.5 t ha-1 yr-1)
0.00
0.36
0.50
0.00
0.36
0.50
0.00
0.36
0.50
0.00
0.36
0.50
0.00
0.36
0.50
0.00
0.36
0.50
N20 efflux (t ha-1 CO2-C equivalent)
CH4 efflux (t ha-1 CO2-C
equivalent)
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
Total C sequestered
Site balance of C sequestration
0.86
0.86
0.86
0.86
0.86
0.86
0.86
0.86
0.86
0.86
0.86
0.86
34
Pig slurry Sewage
sludge
BIOENERGY CROPS AND BIOREMEDIATION
Table 2.14. Carbon equivalent fluxes (C t ha-1) on a Miscanthus plantation four
years after planting, with organic wastes applied after one year. Application rates
the same as Table 2.13.
4th year - Miscanthus
No
addition
Cattle
FYM
Pig FYM
Cattle
slurry
Pig slurry Sewage
sludge
C in litter (6 t ha-1 DM)
C in roots (9 t ha-1 DM)
C in SOM (0.5 t ha-1 yr-1)
C from waste to SOM (10%)
2.70
4.05
2.00
0.00
2.70
4.05
2.00
0.47
2.70
4.05
2.00
0.41
2.70
4.05
2.00
0.22
2.70
4.05
2.00
0.11
2.70
4.05
2.00
0.37
N20 efflux (t ha-1 CO2-C equivalent)
CH4 efflux (t ha-1 CO2-C
equivalent)
0.00
0.00
0.60
0.44
0.60
0.37
0.60
0.20
0.60
0.10
0.60
0.35
Total C sequestered
Site balance of C sequestration
8.75
8.75
9.22
8.19
9.16
8.18
8.97
8.17
8.86
8.16
9.12
8.18
Table 2.15. Carbon equivalent fluxes (C t ha-1) on a Miscanthus plantation after
the full economic rotation of 25 years, with organic wastes applied after one year,
and again every three years. Application rates the same as Table 2.13.
25th year - Miscanthus
C in litter (6 t ha-1 DM)
C in roots (15 t ha-1 DM)
C in SOM (0.5 t ha-1 yr-1)
C from waste to SOM (10%)
N20 efflux (t ha-1 CO2-C equivalent)
CH4 efflux (t ha-1 CO2-C
equivalent)
Total C sequestered
Site balance of C sequestration
No
addition
Cattle
FYM
Pig FYM
Cattle
Slurry
Pig
Slurry
Sewage
Sludge
2.70
6.75
12.50
0.00
2.70
6.75
12.50
3.78
2.70
6.75
12.50
3.24
2.70
6.75
12.50
1.76
2.70
6.75
12.50
0.86
2.70
6.75
12.50
2.99
0.00
0.00
4.80
3.49
4.80
2.99
4.80
1.63
4.80
0.80
4.80
2.76
21.95
21.95
25.73
17.44
25.19
17.40
23.71
17.28
22.81
17.22
24.94
17.38
35
BIOENERGY CROPS AND BIOREMEDIATION
Table 2.16. Carbon equivalent fluxes (C t ha-1) on a Miscanthus plantation four
years after planting, with organic wastes applied every year, after year one.
Application rates the same as Table 2.13.
4th year - Miscanthus
No
addition
Cattle
FYM
Pig FYM
Cattle
slurry
Pig slurry Sewage
sludge
C in litter (6 t ha-1 DM)
C in roots (9 t ha-1 DM)
C in SOM (0.5 t ha-1 yr-1)
C from waste to SOM (10%)
2.70
4.05
2.00
0.00
2.70
4.05
2.00
1.89
2.70
4.05
2.00
1.62
2.70
4.05
2.00
0.88
2.70
4.05
2.00
0.43
2.70
4.05
2.00
1.49
N20 efflux (t ha-1 CO2-C equivalent)
CH4 efflux (t ha-1 CO2-C equivalent)
0.00
0.00
2.40
1.75
2.40
1.50
2.40
0.81
2.40
0.40
2.40
1.38
Total C sequestered
Site balance of C sequestration
8.75
8.75
10.64
6.49
10.37
6.47
9.63
6.42
9.18
6.38
10.24
6.46
Table 2.17. Carbon equivalent fluxes (C t ha-1) on a Miscanthus plantation after
the full economic rotation of 25 years, with organic wastes applied every year,
after year one. Application rates the same as Table 2.13.
25th year - Miscanthus
C in litter (6 t ha-1 DM)
C in roots (15 t ha-1 DM)
C in SOM (0.5 t ha-1 yr-1)
C from waste to SOM (10%)
N20 efflux (t ha-1 CO2-C equivalent)
CH4 efflux (t ha-1 CO2-C equivalent)
Total C sequestered
Site balance of C sequestration
No
addition
Cattle
FYM
Pig FYM
Cattle
slurry
2.70
6.75
12.50
0.00
2.70
6.75
12.50
11.34
2.70
6.75
12.50
9.72
2.70
6.75
12.50
5.29
2.70
6.75
12.50
2.59
2.70
6.75
12.50
8.96
0.00
0.00
14.40
10.48
14.40
8.98
14.40
4.89
14.40
2.40
14.40
8.28
21.95
21.95
33.29
8.41
31.67
8.29
27.24
7.95
24.54
7.75
30.91
8.23
36
Pig slurry Sewage
sludge
BIOENERGY CROPS AND BIOREMEDIATION
2.2.3 References
Aitken, M N (1997). Use of sewage sludge on agricultural land. SAC
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Arcara, P G; Gamba, C; Bidini, D & Marchetti, R (1999). The effect of
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Bender, M & Conrad, R (1995). Effects of CH4 concentrations and soil
conditions on the induction of CH4 oxidation activity. Soil Biology and
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inventory of nitrous oxide emissions from farmed livestock.
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emissions following the land spreading of solid manures. In: Gaseous
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Wallingford, Oxon.. 275-280.
Chambers, B J; Lord, E I; Nicholson, F & Smith, K A (1999a).
Predicting nitrogen availability and losses following application of
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15. 137-143.
Chambers, B J; Nicholson, N; Smith, K; Pain, B; Cumby, T &
Scotford, I (1999b). Managing Livestock Manures. Booklet 1.
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BIOENERGY CROPS AND BIOREMEDIATION
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MAFF,
Chang, C; Cho, C M & Janzen, H H (1998). Nitrous oxide emission
from long-term manured soils. Soil Science Society of America
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permanent grass: seasonal and diurnal fluctuations as influenced by
manuring and fertilisation. Soil Biology and Biochemistry 15. 531536.
Coyne, M S; Villalba, A & Blevins, R L (1995). Nitrous oxide loss from
poultry manure-amended soil after rain. Journal of Environmental
Quality 24. 1091-1096.
Dendooven, L; Splatt, P; Pemberton, E; Ellis, S & Anderson, J M
(1997). Controls over denitrification and its gaseous products in a
permanent pasture soil.
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Wallingford, Oxon.. 19-26.
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atmospheric trace gases. Nature 319. 109-115.
Dobbie, K E; McTaggart, I P & Smith, K A (1999). Nitrous oxide
emissions from intensive agricultural systems: Variations between
crops and seasons, key driving variables, and mean emission factors.
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HMSO, London.
Egginton, G M & Smith, K A (1986). Nitrous oxide emission from a
grassland soil fertilised with slurry and calcium nitrate. Journal of Soil
Science 37. 59-67.
Eichner, M J (1990). Nitrous oxide emissions from fertilised soils:
summary of available data. Journal of Environmental Quality 19.
272-280.
Ellis, S; Goulding, K & Dendooven, L (1997). Effect of pH on nitrous
oxide emissions from grassland soil. In: Gaseous Emissions from
Grasslands (Eds. S C Jarvis & B F Pain). CABI, Wallingford, Oxon..
181-188.
Elowson, S & Rytter, L (1993). Spatial distribution of roots and root
nodules and total biomass production in a grey alder plantation on
sandy soil. Biomass & Bioenergy 5. 127-135.
Ericsson, T; Rytter, L & Vapaavuor, E (1996). Physiology of carbon
allocation in trees. Biomass and Bioenergy 11. 115-127.
Ferm, M; Kasimir-Klemedtsson, A; Weslien, P & Klemedtsson, L
(1999). Emission of NH3 and N2O after spreading of pig slurry by
broadcasting or band spreading. Soil Use and Management 15. 27-33.
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BIOENERGY CROPS AND BIOREMEDIATION
Firestone, M K & Davidson, E A (1989). Microbiological basis of NO
and N2O production and consumption in soil. In: Exchange of trace
gases between
terrestrial ecosystems and the atmosphere (Eds. M O
Andrae & D S Schimel). John Wiley & Sons Ltd., Chichester. 7-21.
Fowler, D; Skiba, U & Hargreaves, K J (1997). Emissions of nitrous
oxide from grasslands. In: Gaseous Emissions from Grasslands (Eds.
S C Jarvis & B F Pain). CABI, Wallingford, Oxon.. 147-164.
Frost, J P; Stevens, R J & Laughlin, R J (1990). Effect of separation and
acidification of cattle slurry on ammonia volatilisation and on the
efficiency of slurry nitrogen for herbage production. Journal of
Agricultural Science 115. 49-56.
Harrison, R & Webb, J (2001). A review of the effect of N fertilizer type
on gaseous emissions. Advances in Agronomy 73. 65-108.
Hogan, K B (1993). Anthropogenic methane emissions in the United
States. Estimates for 1990. Report to Congress, EPA 430-R-93-003.
USEPA.
Houghton, J T; Meira Filho, L G; Callander, B A; Harris, N;
Kattenberg K & Maskell, A (1996). Climate Change 1995: The
Science of Climate Change. Cambridge University Press, Cambridge.
IPCC (1997). Guidelines for National Greenhouse Gas Inventories.
OECD, Paris.
Jarvis, S C; Hatch, D J; Pain, B F & Klarenbeek, J V (1994).
Denitrification and the evolution of nitrous oxide after the application
of cattle slurry to a peat soil. Plant and Soil 166. 231-241.
Jarvis, S C & Pain, B F (1990). Ammonia volatilisation from agricultural
land. Proceedings of the Fertiliser Society 298. The Fertiliser Society,
Peterborough, UK.
Jarvis, S C; Lovell, R D & Panyides, R (1995). Patterns of methane
emission from excreta of grazing animals.
Soil Biology and
Biochemistry 27. 1581-1588.
Jenkinson, D S (1990). The turnover of carbon and nitrogen in soil.
Philosophical Transactions of the Royal Society, London B329. 361368.
Lessard, R; Rochette, P; Gregorich, E G; Pattey, E & Desjardins, R L
(1996). Nitrous oxide fluxes from manure-amended soil under maize.
Journal of Environmental Quality 25. 1371-1377.
Lodman, D W; Branine, M E; Carmean, B R; Zimmermans, P; Ward,
G M & Johnson, D E (1993). Estimates of methane emissions from
manure of US cattle. Chemosphere 26. 189-199.
McTaggart, I P; Douglas, J T; Clayton, H & Smith, K A (1997).
Nitrous oxide emission from slurry and mineral nitrogen fertiliser
applied to grassland. In: Gaseous Emissions from Grasslands (Eds. S
C Jarvis. & B F Pain). CABI, Wallingford, Oxon.. 201-209.
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BIOENERGY CROPS AND BIOREMEDIATION
Maag, M; Lind, A-M & Eiland, F (1996). Emission of nitrous oxide and
denitrification from Danish soils amended with slurry and fertiliser. In:
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Menzi H; Katz P; Frick R; Fahrni M & Keller M (1997). Ammonia
emissions following the application of solid manure to grassland. In:
Gaseous Emissions from Grasslands (Eds. S C Jarvis & B F Pain).
CABI, Wallingford, Oxon.. 265-274.
Mogge, B; Heinemeyer, O; Kaiser, E-A & Munch, J Ch (1996). N2O
emissions of forest soils in northern Germany: seasonal variability and
influencing parameters. In: Progress in Nitrogen Cycling Studies. 8th
Nitrogen Workshop (Eds. O Van Cleemput, G Hofman & A
Vermoesen). Kluwer Academic Publishers, Dordrecht, Netherlands.
585-588.
Mosier, A R; Hutchinson, G L; Sabey, B R & Baxter, J (1982). Nitrous
oxide emissions from barley plots treated with ammonium nitrate or
sewage sludge. Journal of Environmental Quality 11. 78-81.
Paul, J W; Beauchamp, E G & Zhang, X (1993). Nitrous and nitric
oxide emissions during nitrification and denitrification from manureamended soil in the laboratory. Canadian Journal of Soil Science 73.
539-553.
Palm, C A & Rowland, A P (1997). A minimum dataset for
characterization of plant quality for decomposition. In: Driven by
Nature: Plant litter quality and decomposition (Eds. G Cadisch & K
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Petersen, S O (1999). Nitrous oxide emissions from manure and inorganic
fertilisers applied to spring barley. Journal of Environmental Quality,
28. 1610-1618.
Rytter, L; Slapokas’ T & Granhall, U (1989). Woody biomass and litter
production of a fertilised grey alder plantations on a low-humified peat
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Schütz, H; Seiler, W & Renneberg, H (1990). Soil and land use related
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Scott, A; Ball, B C; Crichton I J & Aitken M N (2000). Nitrous oxide
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Management 41. 143-165.
40
BIOENERGY CROPS AND BIOREMEDIATION
Smith, K A & Chambers, B J (1993). Utilising the nitrogen content of
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Smith, S R (1991). Effects of sewage sludge application on soil microbial
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41
BIOENERGY CROPS AND BIOREMEDIATION
2.3
HEAVY METALS AND OTHER CONTAMINANTS
FIONA NICHOLSON
2.3.1 Introduction
Livestock manures and other organic wastes applied to energy crops are an
important source of plant nutrients such as nitrogen (N) and phosphorus
(P), and also supply valuable quantities of organic matter to the soil.
However, they may also contain a range of contaminants that can be
harmful to plant, ecosystem or human health.
Heavy metals
Among the most important contaminants are the heavy metals. Zinc (Zn),
copper (Cu) and nickel (Ni) are essential trace elements for plant growth,
although they can be phytotoxic if present at excessive levels in the soil
(Alloway, 1990). However, chromium (Cr), cadmium (Cd), lead (Pb),
arsenic (As) and mercury (Hg) have no known biological function and can
be harmful if they enter the wider environment. Because heavy metals
cannot be broken down into less harmful by-products, phytoremediation
strategies focus on their accumulation in above-ground plant parts and
subsequent removal from the contaminated site.
Heavy metals may be present in different forms in manures and organic
wastes, which could affect the amount and rate available for plant uptake.
For example, metals in animal manures will be present mainly in soluble
forms or bound to organic matter. In contrast, biosolids or industrial wastes
may contain a large proportion of metals strongly bound to mineral
particles (in particular Fe, Al and Mn oxides), that are much less available
to plants. Once incorporated into the soil, a number of soil properties will
affect plant metal uptake including pH, cation exchange capacity, redox
conditions, organic matter content and clay content.
Organic contaminants
Other contaminants often present in waste materials include a wide range of
organic compounds with the potential to exert a health or environmental
hazard. These include polynuclear aromatic hydrocarbons (PAHs),
polychlorinated biphenyls (PCBs), phthalate acid esters, phenols,
polychlorinated dibenzo-p-dioxins and furans (PCDD/Fs), herbicides and
organochlorine pesticides, which all have different toxicities and
environmental effects. In contrast to heavy metals, organic pollutants can
often be completely broken down by plants into less harmful metabolites.
There is no evidence to suggest that organic contaminants from different
wastes will behave differently once in the soil, however, soil properties
may affect plant uptake and degradation rates.
42
BIOENERGY CROPS AND BIOREMEDIATION
Other contaminants
Manures and wastes may also contain pathogenic micro-organisms which
have the potential to cause plant, animal or human disease. This form of
contamination is not considered further in this review, although it may be
an important issue when considering the application of livestock and other
organic manures to biomass crops.
2.3.2 Concentrations of contaminants in livestock manures
and other wastes
Livestock manures
There is good recent information available on heavy metal concentrations
in livestock manures collected from farms in England and Wales
(Nicholson et al., 1999). This study of 85 manure samples found that the
Zn and Cu concentrations of pig and poultry manures were higher than for
cattle manures due to supplementation of these metals in pig and poultry
diets (Table 2.18). Heavy metal loading rates where manures are applied at
Code of Practice (MAFF, 1998) rates of 250 kg ha-1 total N were calculated
and are also given in Table 2.18.
There is much less comprehensive information available on the levels of
organic contaminants in manures, and very little from the UK. Raszyk et
al., 1998 looked at concentrations of 16 PAHs in feedstuffs, drinking water,
stable dust, pig slurry, road dust and soil on three pig farms and two cattle
farms in the Czech Republic in 1995 and 1996. The average sum of 16
PAHs for pig slurry was 543.2 g kg-1 dry matter, the average sum of seven
carcinogenic PAHs was 47.0 g kg-1 dry matter and the average BaP
concentration was 2.3 mu g kg-1 of dry matter.
Chrysene,
benzo(a)anthracene and benzo(b)fluoranthene were the dominant PAH
carcinogens on pig and cattle farms, whilst among the other PAHs,
phenanthrene, fluoranthene and pyrene were dominant.
In Switzerland, Berset & Holzer, 1995 reviewed the present state of
contamination of agricultural soils and manures (sewage sludge, liquid
manure, and compost) with PAHs and PCBs. Overall PAH concentrations
in cattle slurries were 87-309 g kg-1 (mean 165 g kg-1), pig slurries 66339 g kg-1 (mean 143 g kg-1), sewage sludge 1.7-15 mg kg-1 (mean 6.3
mg kg-1) and compost 0.8-2.7 mg kg-1 (mean 2 mg kg-1). PCB
concentrations were 20 g kg-1 in cattle slurries, 37 g kg-1 in pig slurries
and 32 g kg-1 in compost. A qualitative analysis of the environmental
samples showed that besides the 16 PAHs frequently used for
quantification, mainly alkylated derivatives as well as N-S- and-O-PAHs
were detected.
Some work has also been undertaken looking at the inputs and outputs of
various organic contaminants in dairy cattle. A three month study in the
UK found that the total PCB content (53 congeners) in fresh faeces from
lactating dairy cows averaged 1.3 ( 0.42) g kg-1 (Thomas et al., 1999).
43
BIOENERGY CROPS AND BIOREMEDIATION
Preliminary, unpublished data from Lancaster University (G. Thomas, pers.
comm.) suggest faeces concentrations of 31.2 ng kg-1 furans and 62.5 ng
kg-1 dioxins. Other studies have quantified the percentage of organic
contaminant intake which was measured in manure, although manure
concentrations were not specifically reported (Welsch-Pausch &
McLachlan, 1998; McLachlan & Richter, 1998; Stephens et al., 1995).
From the limited information available, it appears that levels of PCDD/Fs,
PAHs and PCBs are much lower in animal manures than in biosolids. In
addition, animal manures are often stored for relatively long periods prior
to land spreading (6 -12 months) and during this time it would be expected
that some organic contaminants would degrade, and the concentrations
reduced compared to fresh excreta.
Biosolids
Heavy metals may often be present at relatively high levels in biosolids
both from industrial and domestic sources, although in recent years
concentrations have decreased, due mainly to improved trade effluent
controls and the adoption of cleaner manufacturing technologies (Smith,
1996). Inputs of heavy metals to agricultural soils are controlled under the
Sludge (Use in Agriculture) Regulations (SI, 1989), which enforces
provisions of EC Directive 86/278/EEC, by means of sludge concentration
limits, soil limits and maximum permitted application rates.
A survey of biosolids production, treatment, recycling and disposal in the
UK was undertaken for the financial year 1996/7 (Gendebien et al., 1999).
The report includes information on weighted average heavy metal
concentrations in biosolids used in agriculture and from this heavy metal
loading rates can be derived, assuming sludges are applied at rates
equivalent to 250 kg ha-1 total N (Table 2.18).
The range of organic contaminants known to be present in biosolids is
extensive and diverse. For example, 332 different compounds were
identified in German biosolids by Drescher-Kaden et al. (1992).
Concentrations of organic pollutants in biosolids were not measured as part
of the UK sewage sludge survey (Gendebien et al., 1999), as there are
currently no limit values set for biosolids or soils in the UK. However,
Smith (1996) summarised the literature published to date and reported
concentration ranges for some of the principal groups of organic
contaminants. The highest concentrations were for linear alkylbenzene
sulphonates (LAS) at 50-15000 mg kg-1 ds, alkylphenols (100-3000 mg kg-1
ds) and phthalates (1-100 mg kg-1 ds).
44
BIOENERGY CROPS AND BIOREMEDIATION
Table 2.18. Heavy metal concentrations and loading rates from organic manures and other wastes (application rates equivalent to
250 kg ha-1 total nitrogen, except for paper sludge where the industry maximum of 15 t ha -1 was used).
Manure type
Dry
matter
(%)
Total N
(kg t-1
or m-3)
Concentration
Zn
Cu
Ni
Livestock manures1
Cattle FYM
Pig FYM
Dairy slurry
Beef slurry
Pig slurry
Broiler/turkey litter
Layer manure
25
25
10
10
10
60
30
6.0
7.0
4.5
3.5
7.0
29
15
17
60
17
17
65
130
175
4.0
42
4.5
4.5
47
19
27
0.7
1.3
0.6
0.6
1.4
2.4
3.0
Biosolids2
15
6.5
119
85
8.6
Paper sludge3
22
14
13
1.0
1
2
3
0.3
Pb
Cr
(g t-1 or m-3)
0.6
0.8
0.7
0.7
0.8
2.0
2.7
33
0.5
0.5
0.6
0.6
0.6
0.7
1.2
1.7
24
1.5
Source : Nicholson et al., 1999
Source : Gendebien et al., 1999
Source : Davis & Rudd, 1998
45
Loading rate
As
Cd
Zn
Cu
Ni
Pb
Cr
(kg ha-1)
As
Cd
0.3
0.2
0.2
0.2
0.2
0.3
0.1
0.06
0.06
0.03
0.03
0.04
0.33
0.39
0.7
2.1
0.9
1.2
2.3
1.1
2.9
0.2
1.5
0.3
0.3
1.7
0.2
0.5
0.03
0.05
0.03
0.04
0.05
0.02
0.05
0.03
0.03
0.04
0.05
0.03
0.02
0.05
0.02
0.02
0.03
0.04
0.02
0.01
0.03
0.01
0.01
0.01
0.02
0.01
<0.01
<0.01
0.003
0.002
0.002
0.002
0.001
0.003
0.007
0.9
0.50
4.6
3.3
0.33
1.27
0.92
0.03
0.019
-
<0.25
4.7
0.8
0.06
0.03
0.17
-
0.005
BIOENERGY CROPS AND BIOREMEDIATION
Industrial wastes
Approximately 6.8 million tonnes (fresh weight) of industrial wastes are recycled
to land in England and Wales each year (Gendebien et al., 2001). Various types of
wastes are applied including a range of vegetable and animal wastes (5.05 Kt),
food industry waste (1,409 Kt), paper sludge (322 Kt), rocks, subsoils and
contaminated soil/subsoil (15 Kt), cement waste (12.2 Kt), other biological
treatment plant sludges (4.3 Kt), textile waste (3.5 Kt) and small quantities of
leather and tannery wastes, fly ash and other mineral wastes (Gendebien et al.,
2001).
Data on the composition of selected industrial wastes are reported by Gendebien et
al. (2001). These were derived from a limited number of sources in the UK over a
period of ten years and may not be entirely representative, although they do
provide a broad indication of the composition of the wastes. As an example, heavy
metal application rates for paper sludge (Table 2.18) were calculated using the
median heavy metal analysis of 25 samples reported by Davis & Rudd (1998) and
assuming the industry maximum application rate of 15 t ha-1. Note that this data
should be treated with some caution, due to the variable nature of paper sludge
from different mills and production processes.
There is little information on other contaminants in industrial wastes. Paper
sludges are known to contain dioxins and organohalogens, especially where
chlorine bleaching was used, and wastes from the textile and tanning industries
may contain organic dyes and pesticide residues (Anon, 1998).
Summary
Assuming an application rate equivalent to 250 kg ha-1 total N, biosolids will be a
more important source of heavy metals and organic contaminants to biomass crops
than livestock manures. Whilst the risks of soil metal contamination from
biosolids applications are recognised and regulated for, there is no current control
over inputs of heavy metals from livestock manures and care should be taken to
ensure that soil concentrations do not become excessively high, especially where
pig and poultry manures are applied. Industrial wastes may supply large amounts
of certain metals (e.g. Zn from de-inked paper sludge and Cr from tanning industry
wastes). The scientific consensus is that, based on the limited information
available, organic contaminants in biosolids applied to agricultural land are
unlikely to cause significant environmental or human health problems (Smith,
1996).
46
BIOENERGY CROPS AND BIOREMEDIATION
2.3.3 Contaminant uptake/removal by biomass crops
Heavy metals
Phytoremediation strategies for heavy metal contaminated sites have tended to
concentrate on the use of metal tolerant or hyper-accumulator species (e.g.
Brassica spp., Thlaspi caerulescens) which usually produce relatively small
quantities of biomass. However, high biomass producers such as willow (Salix
spp.) have been increasingly used as potential phytoremediator crops.
In a Swedish study, high Cd concentrations were found in willow shoots and the
authors concluded that Salix viminalis could potentially remove significant
amounts of Cd from the soil, although it would take 12-20 years to return a soil
with 600 g ha-1 (c. 0.15 mg kg-1) Cd to half its original value (Eriksson & Ledin,
1999). Nielsen (1994), also found that willow takes up large amounts of Cd and
Zn, but Labrecque et al., 1995) reported that Salix discolor and Salix viminalis
were less able to take up Ni, Hg, Cu and Pb.
There are large differences in the efficiency of metal uptake between different
biomass crops, plant species and clones. The EU-funded BIORENEW project is
investigating bioremediation of contaminated land using biomass crops and is
currently assessing the metal tolerance of different willow varieties (Watson et al.,
1999). Other European work has looked at acclimation of Salix species to metal
(Zn, Cu and Cd) stress in an attempt to induce increased resistance (Punshon &
Dickinson, 1997), and at differences in tolerance and metal accumulation between
different Salix clones (Landberg et al., 1996). Similarly, work in the USA has
looked at differences in metal uptake between Salix clones grown hydroponically
(Punshon & Dickinson, 1999), whilst Rugh et al., (1998) reported initial work on
the potential of transgenic yellow poplar trees to remediate Hg polluted soils.
Reported concentrations of metals in biomass crops are given in Table 2.19,
although often these are from plants grown on contaminated or heavily sludged
soils. There can be large differences in concentrations between different plant
parts. For example, Punshon & Dickinson (1997) reported the magnitude of Cu
accumulation in the order roots > wood > new stem > leaves; whereas for Cd the
order was leaves > stem > wood > roots. Roots, wood and stems can immobilise
metals for a number of years, as opposed to leaves which are shed annually
returning any metals they contain to the soil. It is most useful in terms of
remediation potential for metals to accumulate in the harvested plant parts (i.e.
wood, stems and leaves). Note that the time of sampling can also influence the
heavy metal concentration in the leaves (Riddel-Black, 1994) as can interactions
with other soil heavy metals (Punshon & Dickinson, 1997).
47
BIOENERGY CROPS AND BIOREMEDIATION
Table 2.19. Heavy metal concentrations in biomass crops and calculated metal uptake rates.
Metal
Crop
Concentration
Comments
Source
Metal contaminated soils
Metal contaminated soils
8-30 year old plantations
8-30 year old plantations
Field contaminated with Cd
Heavy sludge applications
Plants grown in culture solutions
Sludge applications made
Borjesson (1999)
Felix et al. (1999)
Dickinson (1997)
Dickinson (1997)
Eriksson & Ledin (1999)
Eriksson & Ledin (1999)
Felix (1997)
Riddell-Black (1994)
Punshon & Dickinson (1997)
Labrecque et al. (1995)
(mg kg-1 DM)
Cd
Salix - shoots
Salix viminalis - aerial tissues
Salix -foliage
Salix - wood
Salix viminalis - stem-wood
Salix viminalis - leaves
Salix viminalis - aerial tissues
Salix (4 spp) - stems
Salix (4 spp) -aerial tissues
Salix (2 spp) - leaves
0.4 - 3.9
22
43.9
76.4
0.35 - 2.43
0.31 - 1.96
8.3-22.2
3.3-7.7
<100
<1.8
Cu
Salix - foliage
Salix - wood
Salix caprea - foliage and woody branches
Salix (7 spp) - foliage and woody branches
Salix (4 spp)-stems
Salix (4 spp) -aerial tissues
Salix (2 spp) - leaves
4.2
4.6
681
183-509
5.9- 8.2
25-100
<20.7
Metal contaminated soils
Metal contaminated soils
3-year old trees on highly contaminated soil
1-year old trees on highly contaminated soil
Heavy sludge applications
Plants grown in culture solutions
Sludge applications made
Dickinson (1997)
Dickinson (1997)
Dickinson (2000)
Dickinson (2000)
Riddell-Black (1994)
Punshon & Dickinson (1997)
Labrecque et al. (1995)
Zn
Salix - foliage
Salix - wood
Salix (3 spp) -roots
Salix (3 spp) - stems
Salix (4 spp)-stems
Salix (2 spp) - leaves
87.1
77.3
<50
<8
95-156
<560
Metal contaminated soils
Metal contaminated soils
Plants grown on mining spoil
Plants grown on mining spoil
Heavy sludge applications
Sludge applications made
Dickinson (1997)
Dickinson (1997)
Dickinson et al. (1994)
Dickinson et al. (1994)
Riddell-Black (1994)
Labrecque et al. (1995)
Pb
Salix - foliage
Salix - wood
Salix (2 spp) - leaves
17.3
157.4
<6.5
Metal contaminated soils
Metal contaminated soils
Sludge applications made
Dickinson (1997)
Dickinson (1997)
Labrecque et al. (1995)
Ni
Salix (4 spp)-stems
Salix (2 spp) - leaves
0.9-1.4
<20
Heavy sludge applications
Sludge applications made
Riddell-Black (1994)
Labrecque et al. (1995)
Hg
Salix (2 spp) - leaves
<20
Sludge applications made
Labrecque et al. (1995)
48
BIOENERGY CROPS AND BIOREMEDIATION
Soil factors such as soil pH and clay content can strongly influence plant heavy
metal uptake. Some researchers have reported manipulating soil conditions or
using soil amendments (e.g. chelating agents) to increase metal uptake. For
example, Kayser et al. (2000) investigated the efficiency of several crops
(including Salix viminalis) at phyto-extracting Zn, Cd and Cu from calcareous soils
and found that plant accumulation of these metals increased by a factor of 2-3
where nitrilotriacetate (NTA) and elemental sulphur were applied to the soil.
However, the use of soil amendments to increase metal solubility can also increase
the risks of leaching or downward migration of the metals. It is also important to
note that metals may be present in wastes and soils in different forms, and that
some metal fractions are so strongly bound to mineral particles that they may never
be taken up by plants and removed from the soil.
There has been relatively little work published specifically looking at metal uptake
following organic manure applications to biomass crops. However, metal uptake
rates from a study where sewage sludge was applied to Salix crops at 12.5 t ds ha-1
yr-1 for 6 years (Hasselgren, 1999) are reported in Table 2.20. Typical UK sludge
and poultry manure metal application rates can be compared with the stem uptake
rates reported by Hasselgren (1999) and indicate that all metals would be added at
higher rates than they could be removed by the crop, leading to a net accumulation
in the soils (Table 2.20).
Table 2.20. Salix stem uptake of sludge heavy metals, amounts of metals
supplied by UK sludges and poultry manure applied at 250 kg ha -1 total N
yr-1, and calculated metal accumulation rates.
Heavy
metal
Zn
Cu
Pb
Cd
Ni
Cr
1
2
3
Average stem
Amount of metal
metal uptake
supplied by UK
rate (g ha-1 yr-1)1 sludge (g ha-1 yr-1)
600
59
30
11
8.2
4.8
Net metal
accumulation
(g ha-1 yr-1)
4562
328
1273
19
328
904
3962
269
1243
8
320
899
Amount of metal
supplied by poultry
manure2 (g ha-1 yr-1)
Net metal
accumulation
(g ha-1 yr-1)3
2900
500
50
7
50
30
2300
441
20
4
42
25
Source : Hasselgren (1999).
Laying hen manure (see Table 2.18).
Assumes metals applied in poultry manure are taken up at the same rate as those in sludge.
Labrecque et al. (1995) calculated the transfer coefficients (TC) of metals from
soils to two Salix species where dried, pelleted sludge was applied. Cd and Zn
were the most readily transferred metals, with 15-25% of added sludge metals
taken up into the biomass, followed by Ni, Hg, Cu and Pb. Sludge applications
increased levels of metals in the soil proportionately with the amounts applied in
the sludge. However, this did not necessarily translate into an increased plant
concentration. The authors found there was a good relationship between soil and
plant concentrations of Cd and Zn, while plant concentrations of Cu, Hg, Ni and Pb
were less dependent on the soil concentrations.
49
BIOENERGY CROPS AND BIOREMEDIATION
In a later study, Labrecque et al. (1998) reported the differences between the
amounts of metal introduced into the soil with dried, pelleted sludge applied at
different rates (100, 200 and 300 kg ha-1 of available N) and those removed
following biomass harvest. The net accumulation of Zn in the soil after two years
of sludge applications ranged from c. 2.5 to c. 8.4 kg ha-1 and for Cu from c. 1.4 to
c. 4.4 kg ha-1 (Table 2.21). The amount of Zn removed by the crop increased
slightly with increased sludge application rate, whilst the amount of Cu removed
was similar at all application rates. Analysis of soil (0-20 cm depth) confirmed
that soil Zn and Cu levels had generally increased compared with the pre-planting
levels; however levels of soil Ni, Cr and Cd were largely unaffected by the two
sludge applications.
Table 2.21. Net soil accumulation of metals after 2 years of sludge
applications.
Metal
Sludge
content
(mg kg-1 ds)
Amount of
metal applied
(kg ha-1)
Amount of metal
removed in
biomass (kg ha-1)1
Net soil metal
accumulation
(kg ha-1)2
Zinc
404
3.03
6.06
9.09
0.57
0.65
0.72
2.46
5.41
8.37
Copper
201
1.51
3.02
4.53
0.07
0.11
0.09
1.44
2.91
4.44
Source: Labrecque et al., 1998
1
Calculated by difference
2
Mean results of 2 sites, 2 species and 2 planting densities
These results indicate that Salix crops can remove heavy metals from soils,
although the removal rate depends on the metal concerned, soil properties, the
plant species/variety grown and the amount of biomass produced. However, where
organic manure applications are made, metals will be added at greater rates than
they can be removed by a biomass crop such as Salix, leading to an accumulation
in the soil. Assuming a typical UK biosolids application to Salix at a rate
equivalent to 250 kg N yr-1 would lead to a net soil accumulation of c. 4 kg Zn ha-1
yr-1 (Table 2.20), and assuming that the soil currently has the UK median Zn
concentration of 82 mg kg-1 (McGrath & Loveland, 1992), then it would take
around 115 years for the soil to reach the limit concentration of 200 mg kg -1 (DoE,
1996). Similarly, for poultry manure applications it would take around 160 years
to reach the soil limit concentration for Zn.
50
BIOENERGY CROPS AND BIOREMEDIATION
Other contaminants
Hybrid poplar trees have been used to uptake, hydrolyse and dealkylate the
pesticide atrazine to less toxic metabolites (Burken & JL, 1997; Burken &
Schnoor, 1996). Results indicated that poplar cuttings were able to take up the
majority of the applied atrazine that was not strongly sorbed to the soil organic
fraction. Hybrid poplar cuttings have also been used to uptake and translocate 1,4dioxane (a suspected carcinogen) to leaf surfaces (Aitchison et al., 2000). Around
80% of the dioxane taken up by poplars was transpired to the atmosphere where it
can be readily dispersed and photo-degraded. Other researchers in the USA
(Orchard et al., 2000) have used poplars to take up the high explosive 2,4,6trinitrotoluene (TNT) and the carcinogenic degreaser trichloroethylene (TCE).
This work has looked exclusively at the remediation of contaminated sites and has
not addressed the use of these plants to take up contaminants from organic manures
or other wastes.
Ultimate fate of metals and other contaminants in biomass crops
At harvest, the above-ground parts of biomass crops will be removed from the
growing site and transported to the power station. During combustion all the
heavy metals, except Hg, will be concentrated in the resultant ash. Because Hg is a
volatile metal, during combustion it will be released in the gaseous state, and
eventually be oxidised and scavenged by wet or dry deposition processes at various
distances from the source (Steinnes, 1990). Some researchers have found elevated
levels of Hg in soils near crematoria caused by the volatilisation of Hg in dental
fillings during cremation, which was subsequently re-deposited to the nearby soil
(Nieschmidt & Kim, 1997). A similar situation could occur near power stations
where biomass crops grown on Hg contaminated soils are burnt.
Metals on particulates in the flue gases may be discharged to the atmosphere, via
the stack, and subsequently re-deposited onto land. However, appropriate
technology to improve the quality of stack emissions should minimise this
problem, although there may be considerable expense involved. Metals in the
bottom and fly ash could potentially be recovered for re-use, however this is likely
to be very expensive and the quantities of ash generated are unlikely to make
recovery an economic proposition. It is more likely that the ash will be buried in
landfill sites where conditions may be such that metal containing leachates could
enter the environment.
2.3.4 Other pathways of contaminant movement and environmental
effects
Hasselgren (1999) reported that most of the metals applied to Salix crops in sewage
sludge remained in the upper 10 cm soil layer, with total soil Zn and Cu
concentrations increasing by 17 and 4 times, respectively, where sludge was
applied at 12.5 t ds ha-1 yr-1 for 6 years. Hasselgren (1998) also found that where
sewage sludge was applied to Salix at rates up to 20 t ha-1 ds, the groundwater (at
1-2.5 m) had double the concentration of Cu, Pb and Hg, and 3-5 times the
51
BIOENERGY CROPS AND BIOREMEDIATION
concentration of Zn and Ni after sludge applications, although Cd concentrations
were below detection limits.
Most previous research indicates that heavy metals applied in biosolids would
accumulate in the upper soil layers, with relatively little lost via leaching (Smith,
1996). Nevertheless, there could be a greater risk of leaching if soil amendments
are used to increase plant metal uptake.
Salix leaves have been shown to accumulate high concentrations of some metals
and a further risk to the environment could come where animals consume these
leaves and accumulate high levels of metals in their body tissues.
2.3.5 Conclusions
1.
Biomass crops can remove heavy metals and organic contaminants from soils,
although the removal rate depends on contaminant and soil properties, the
plant species/variety grown and the amount of biomass produced.
2.
Where organic manure (livestock manure, biosolids, industrial waste)
applications are made, heavy metals will generally be added at greater rates
than they can be removed by a biomass crop, leading to an accumulation in
the soil.
3.
Very little is known about the types and quantities of organic contaminants in
organic manures or how quickly they could be removed or broken down by
biomass crops.
4.
After combustion, if the ash is buried in landfill sites, conditions may be such
that heavy metal leachates could enter the environment.
2.3.6 References
Aitchison, E; Kelley, S; Alvarez, P & Schnoor, J (2000). Phytoremediation of
1,4,-dioxane by hybrid poplar trees. Water Environment Research 72 (3).
313-321.
Alloway, B J (1990). Heavy Metals in Soils. Blackie, London.
Anon. (1998). Clampdown on the horizon for land spreading of industrial wastes.
ENDS Report 281. 27-29.
Berset, J & Holzer, R (1995). Organic micropollutants in Swiss agriculture distribution of polynuclear aromatic-hydrocarbons (PAH) and polychlorinatedbiphenyls (PCB) in soil, liquid manure, sewage-sludge and compost samples a
comparative-study.
International Journal of Environmental Analytical
Chemistry 59. 145-165.
Borjesson, P (1999). Environmental effects of energy crop cultivation in Sweden
- I. Identification and quantification. Biomass and Bioenergy 16. 137-154.
52
BIOENERGY CROPS AND BIOREMEDIATION
Burken, J & Schnoor, J (1996). Phytoremediation: plant uptake of atrazine and
role of root exudates. Journal of Environmental Engineering 122 (11). 958963.
Burken, J G & Scnoor, J L (1997). Uptake and metabolism of atrazine by poplar
trees. Environmental Science and Technology 31 (5). 1399-1406.
Davis, R D & Rudd, C (1998). Investigation of the criteria for and guidance on
the landspreading of industrial wastes. Final report to the Environment
Agency. WRc Report 4088/7. Environment Agency R&D Technical Report
P193.
Dickinson, N (2000). Strategies for sustainable woodland on contaminated soils.
Chemosphere 41. 259-263.
Dickinson, N M (1997). Rehabilitation and remediation of metal-contaminated
soils using trees. 4th International Conference on the Biogeochemistry of
Trace Elements, Berkeley, California. 437-8.
Dickinson, N M; Punshon, T; Hodkinson, R B & Lepp, N W (1994). Metal
tolerance and accumulation in willow. Willow Vegetation Filters for Municipal
Wastewaters and Solids, Uppsala, Sweden. 121-127.
DoE (1996). Code of Practice for Agriculture Use of Sewage Sludge. Department
of the Environment.
Drescher-Kaden, U; Bruggeman, R; Matthes, B & Matthies, M (1992).
Contents of organic pollutants in German sewage sludges. In: Effects of
Organic Contaminants in Sewage Sludge on Soil Fertility, Plants and Animals
(Eds. J Hall, D Suauerbeck & P L'Hermite). Commission of the European
Communities, Luxembourg. 14-34.
Eriksson, J & Ledin, S (1999). Changes in phytoavailability and concentration of
cadmium in soil following long term Salix cropping. Water, Air, Soil
Pollution 11. 171-184.
Felix, H (1997). Field trials for in situ decontamination of heavy metal polluted
soils using crops of metal-accumulating plants.
Zeitschrift fuer
Pflanzenernahr und Bodenkunde 160. 525-529.
Felix, H R; Kayser, A & Schulin, R (1999). Phytoremediation, field trials in the
years 1993-1998. 5th International Conference on the Biogeochemistry of
Trace Elements, Vienna, Austria. 8-9.
Gendebien, A; Carlton-Smith, C; Izzo, M & Hall, J (1999). UK Sewage Sludge
Survey: National Presentation. R&D Technical Report P165. Environment
Agency Bristol.
Gendebien, A et al. (2001). Survey of Wastes Spread on Land. Final report to the
European Commission - Directorate General for Environment. Report No.
CO4953-2.
Hasselgren, K (1998). Use of municipal waste products in energy forestry:
highlights from 15 years of experience. Biomass and Bioenergy 15 (1). 7174.
Hasselgren, K (1999). Utilization of sewage sludge in short-rotation energy
forestry : a pilot study. Waste Management and Research 17. 251-262.
53
BIOENERGY CROPS AND BIOREMEDIATION
Kayser, A; Wenger, K; Keller, A; Attinger, W; Felix, H; Gupta, S & Schulin,
R (2000). Enhancement of phytoextraction of Zn, Cd and Cu from calcareous
soil: the use of NTA and sulfur amendments. Environmental Science and
Technology 34 (9). 1778-1783.
Labrecque, M; Teodorescu, T & Daigle, S (1995). Effect of waste water sludge
on growth and heavy metal bioaccumulation of two Salix species. Plant and
Soil 171. 303-316.
Labrecque, M; Teodorescu, T & Daigle, S (1998). Early performance and
nutrition of two willow species in short-rotation intensive culture fertilized
with wastewater sludge and impact on the soil characteristics. Canadian
Journal of Forestry Research 28. 1621-1635.
Landberg, T; Greger, M; Fuge, R; Billet, M & Selinus, O (1996). Differences
in uptake and tolerance to heavy metals in Salix from unpolluted and polluted
areas. Environmental Geochemistry 11 (1-2). 175-180.
MAFF (1998). Code of Good Agricultural Practice for the Protection of Water.
MAFF Publications.
McGrath, S P & Loveland, P J (1992). The Soil Geochemical Atlas of England
and Wales. Blackie Academic and Professional, London.
McLachlan, M & Richter, W (1998). Uptake and transfer of PCDD/Fs by cattle
fed naturally contaminated feedstuffs and feed contaminated as a result of
sewage sludge application. 1. Lactating cows. J Agric. Food Chem. 46.
1166-1172.
Nicholson, F A; Chambers, B J; Williams, J W & Unwin, R J (1999). Heavy
metal contents of livestock feeds and animal manures in England and Wales.
Bioresource Technology 70. 23-31.
Nielsen, K (1994). Environmental aspects of using waste waters and sludges in
energy forest cultivation. Biomass and Bioenergy 6 (1-2). 123-132.
Nieschmidt, A & Kim, N (1997). Effects of mercury release from amalgam
dental restorations during cremation on soil mercury levels of three New
Zealand crematoria. Bulletin of Environmental Contamination and Toxicology
58 (5). 744-751.
Orchard, B J; Doucette, W J; Chard, J K & Bugbee, B (2000). Uptake of
trichloroethylene by hybrid poplar trees grown hydroponically in flow-through
plant growth chambers. Environmental Toxicology and Chemistry 19 (4).
895-903.
Punshon, T & Dickinson, N M (1997). Acclimation of Salix to metal stress. New
Phytologist 137 (2). 303-314.
Punshon, T & Dickinson, N (1999). Heavy metal resistance and accumulation
characteristics in willows. International Journal of Phytoremediation 1 (4).
361-385.
Raszyk, J; Ulrich, R; Gajduskova, V; Salava, J & Palac, J (1998). Occurrence
of carcinogenic polycyclic aromatic hydrocarbons (PAH) on pig and cattle
farms. Veterinarni Medicina 43. 17-25.
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BIOENERGY CROPS AND BIOREMEDIATION
Riddel-Black, D (1994). Heavy metal uptake by fast growing willow species.
Willow Vegetation Filters for Municipal Wastewaters and Solids, Uppsala,
Sweden. 145-152.
Rugh, C; Senecoff, J; Meagher, R & Merkle, S (1998). Development of
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SI (1989). The Sludge (Use in Agriculture) Regulations. SI 1989, No. 1263.
Smith, S. 1996. Agricultural Recycling of Sewage Sludge and the Environment.
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Stephens, R D; Petreas, M X & Hayward, D G (1995). Biotransfer and
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Environmental Pollution 102, 129-137.
55
BIOENERGY CROPS AND BIOREMEDIATION
2.4
IMPACTS ON BIODIVERSITY
CHRIS BRITT
2.4.1 Introduction
Although there has been a vast amount of research on the environmental impacts of
applying sewage sludge, farmyard manures and slurries to agricultural land and
forestry, this has been heavily biased towards studies of impacts on water, soil and
air quality – with relatively little work on the effects on biodiversity. There has
been almost no research on the ecological impacts of applying various waste
materials to short rotation coppice or energy grasses.
Because of this paucity of directly relevant information, provisional assumptions
must be based on an interpretation of indirectly relevant experience (e.g.
biodiversity impacts of waste applications to agricultural or forest crops), within
the context of conditions prevailing within ‘typical’ energy crops. Such an
interpretation firstly requires an overview of energy crops as habitats for wildlife.
Ecology of bioenergy crops
The ecology of short rotation willow and poplar crops in the UK, Sweden and
elsewhere has been quite extensively studied. Although biodiversity in any single
plantation is heavily influenced by factors such as intensity of management,
adjacent land uses and proximity to semi-natural woodland, previous research has
demonstrated the potential value of SRC as a habitat for certain groups of
invertebrates and birds. These include numerous phytophagous insect species,
including pest species such as the blue willow beetle (Phratora vulgatissima)
(Sage & Tucker, 1997 & 1998a), and songbirds (Sage & Robertson, 1996; Coates
& Say, 1999). In contrast, there has been very little research on the ecology of
Miscanthus or other grasses grown as bioenergy crops.
Vegetation
Surveys in 1993 and 1996 have provided quite a comprehensive picture of the
ground flora composition in SRC plantations (Sage & Tucker, 1998b). Over the
two surveys, the most frequently recorded National Vegetation Classification
(NVC) types were various weed and tall herb communities – with an increase in
woodland and scrub communities in 1996. The most frequently recorded species
was the common nettle (Urtica dioica), which occurred in 81% of plots in both
years. Rosebay willowherb (Chamaenerion angustifolium) declined sharply
between the two surveys. In total, 151 plant species were recorded, with 19 of
these present in at least 25% of all plots in one or both years. The mean numbers
of plant species recorded per plot, in five 10 m2 quadrats, were 13.5 in 1993 and
13.8 in 1996. The total number of species recorded per plot ranged from 3-24 in
1993 and 4-31 in 1996.
A detailed study of five SRC sites in southern England gave similar results (Coates
& Say, 1999). Although the total numbers of plant species recorded were
generally quite high (up to 120 at a single site), possibly due to inadequate weed
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BIOENERGY CROPS AND BIOREMEDIATION
control at some sites and the close proximity of semi-natural habitats at all, certain
general trends were recognised in vegetation succession. For example:
 Small annual ruderal species, such as scarlet pimpernel (Anagallis arvensis),
groundsel (Senecio vulgaris) and common chickweed (Stellaria media), were
present at low levels during the early stages of the cropping cycle, and then
tended to disappear.
 Larger ruderals, such as common orache (Atriplex patula), cleavers (Galium
aparine), knotgrass
(Polygonum
aviculare), rosebay willowherb
(Chamaenerion angustifolium), thistles (Cirsium spp.) and docks (Rumex spp.),
occurred at high levels early in the cropping cycle and persisted at reduced
levels.
 Semi-ruderal grasses with a degree of shade tolerance, such as creeping bent
(Agrostis stolonifera), cocksfoot (Dactylis glomerata) and Yorkshire fog
(Holcus lanatus), persisted into the cropping cycle.
 Bramble (Rubus fruticosus agg.) invaded coppice crops later in the cycle.
In SRC where weed control was quite good, and the crop canopy was rapidly
closed, ground vegetation from the third year after planting onwards provided
approximately 0-15% cover. Coates & Say (1999) described the vegetation type
dominating most plots at this time as “mixed semi-ruderal.”
The overall picture, therefore, is of a ground flora that is either very sparse –
because of effective weed control measures and/or heavy shading from the SRC
crop – or of limited diversity and dominated by species of low conservation value,
typical of agricultural weed communities or disturbed land. There are occasional
exceptions of course, particularly where herbicide inputs have been negligible or
nil, and where there are nearby seed sources of woodland species – but the
requirements of good SRC husbandry and the relative scarcity of semi-natural
woodland will always ensure that such sites are atypical.
Soil biodiversity
There is some evidence that willow short rotation coppice plantations can support
large earthworm populations, and that earthworms play a significant role in the
decomposition of leaf litter (Šlapokas & Granhall, 1991). Research in Sweden has
also demonstrated differences between willow species in the rates of leaf
decomposition, with Salix daphnoides seemingly more attractive than S. fragilis or
S. viminalis to all soil organisms – including earthworms (Šlapokas & Granhall,
1991).
Ground-dwelling invertebrates
In an upland (almost 300 m altitude), willow SRC plantation in mid-Wales,
ground-dwelling invertebrates, captured in pitfall traps during the first and second
growing seasons, were fairly typical of the previous land-use (Slater et al., 1997).
The study site was weedy, however, and the authors suggest that in “well
managed” SRC plantations pitfall traps are likely to capture few herbivorous
species and more scavengers and carnivores.
Canopy invertebrates
The large leaf area within a fast-growing willow or poplar SRC crop provides an
important food resource for a wide range of herbivorous invertebrates, which can
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BIOENERGY CROPS AND BIOREMEDIATION
sometimes occur at very high densities. These herbivorous, or phytophagous,
species can, in turn, support a diversity of predatory and parasitic invertebrate
species; as well as numerous insectivorous birds and mammals.
Reddersen (2001) conducted a survey of commercial willow (Salix viminalis) SRC
plantations in Denmark, to quantify catkin abundance. He concluded that the
abundance of flowers in years 2-4 after harvest (willow crops did not flower in the
year after harvest), although variable, was sufficiently high to make willow SRC a
potentially important source of nectar for bees and other nectar-feeding insects.
Birds
Short rotation coppice provides a good habitat for many insectivorous, woodland
bird species – including thrushes (Turdidae), tits (Paridae) and warblers
(Sylviidae). Warblers particularly favour SRC, probably because of the abundant
invertebrate prey that this habitat supports and because it is structurally similar to
early stages of regenerating traditional coppice and to willow carr – two habitats
for which these species are known to have a preference. For example, Fuller &
Green (1998), studying bird populations in adjacent stands of recent, old and
thinned/singled small-leaved lime (Tilia cordata) coppice, found that willow
warbler (Phylloscopus trochilus), garden warbler (Silvia borin) and chiffchaff
(Phylloscopus collybita) were all far more abundant in the recent coppice. Thrush
(Turdidae) densities also tended to be higher in recent coppice. Chaffinches,
however, were least frequently recorded in the recent coppice.
Willow SRC has also been shown to provide a good habitat for pheasants
(Phasianus colchicus).
Mammals
There is little published information on mammal populations in SRC or energy
grass plantations. Among the mammal species that are known to use SRC crops,
rabbits (Oryctolagus cuniculus) and roe deer (Capreolus capreolus) are frequently
common (Sage & Tucker, 1998b). A survey of five SRC sites in southern
England, by Coates & Say (1999), recorded 13 mammal species. Four species
were thought to be common – rabbit, brown hare (Lepus capensis), mole (Talpa
europaeus) and wood mouse (Sylvaemus sylvaticus), although this assumption is
based purely on casual observations for the first three (i.e. only small mammals,
including wood mice, were formally recorded). The other nine species recorded in
SRC were hedgehog (Erinaceus europaeus), common shrew (Sorex araneus),
serotine (Vespertilio serotinus) and pipistrelle (Pipistrellus pipistrellus – both subspecies) bats, bank vole (Clethrionomys glareolus), harvest mouse (Micromys
minutus), brown rat (Rattus norvegicus), badger (Meles meles) and roe deer.
Other species likely to utilise SRC plantations would include predatory species
such as the fox (Vulpes vulpes), stoat (Mustela erminea), weasel (Mustela nivalis)
and polecat (Putorius putorius).
Slater et al. (1997) compared small mammal populations in newly established
willow SRC (during the first year only), and in a nearby rough grassland pasture, at
an upland site (almost 300 m) in mid-Wales. They recorded higher numbers of
wood mice in SRC and higher numbers of field voles (Microtus agrestis) in the
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BIOENERGY CROPS AND BIOREMEDIATION
pasture. Destruction of the ground flora, during SRC crop establishment, resulted
in the initial loss of field voles. Consequently no field voles were trapped in the
first spring after planting, but small numbers were recorded in the autumn. It is
unlikely that SRC plantations with high standards of weed control will support
large vole populations. Almost weed-free SRC crops will also provide relatively
poor habitats for other small mammal groups, such as mice and shrews – due to a
lack of low ground cover (absence of nesting sites and exposure to predators) and
inadequate food resources (e.g. invertebrates, seeds and fruits). Indeed, an
unpublished PhD project by Bodnor (cited by Sage & Tucker, 1998b) found that
both abundance and diversity of small mammals, such as wood mice, were greater
in weedy SRC plots.
Bats make disproportionate use of rivers and lakes for hunting (Vaughan et al.,
1997), but the frequently large populations of arboreal invertebrates in SRC may
make this a suitable habitat for foraging woodland bat species such as the 45 kHz
pipistrelle (Pipistrellus pipistrellus), Myotis spp. and, possibly, horseshoe bats
(Rhinolophus spp.) and Plecotus spp. Further research is needed to confirm the
extent of usage of energy crops by bat species, and to evaluate the potential role of
energy crops in bat conservation.
Effects of waste applications
The application of waste materials to energy crops might be expected to have
significant effects on the flora and fauna within bioenergy crops, such as SRC and
Miscanthus. However, an extensive review of the literature revealed no evidence
of research that looked directly at the ecological effects of applying farm, urban or
industrial waste products to energy crops. There has also been only a limited
amount of research into the ecological effects of applying such materials to other
types of vegetation. Consequently, although relevant research will be referred to
when possible, much of this section must inevitably rely largely upon ‘educated
supposition’. For example, it can safely be assumed that thick applications (or
patchy distribution) of some materials, particularly wastes that are relatively slow
to decompose, will suppress ground flora development. Although this ‘mulching’
of weeds will be beneficial to the crop, it will have generally negative impacts on
biodiversity. On the other hand, increased soil organic matter will be likely to
enhance populations of soil micro-organisms and invertebrates; and applications of
organic manures can provide an important food resource for coprophagous
organisms (e.g. species of Basidiomycete fungi and scarab beetles) – which, in
turn, will be a source of food for species higher in the food chain. Manures may
also be the source of weed seeds. The supply of N in manures, slurries or sewage
sludge will increase crop growth rates, producing more and ‘softer’ foliage with
greater susceptibility to grazing insects.
It is important, of course, to consider all possible direct and indirect ecological
effects of waste applications to bioenergy crops within the context of standard
management practices for these crops. Neither the effects of weed suppression nor
nutrient enrichment referred to above are likely to have particular ecological
significance, if the alternative is increased inputs of herbicides and inorganic N.
Indeed, it could be that organic waste utilisation could provide a net positive effect
in these circumstances.
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BIOENERGY CROPS AND BIOREMEDIATION
2.4.2 Agricultural and municipal wastes
Vegetation
As has been clearly shown previously in this review, there is plentiful evidence that
applying organic ‘waste’ materials such as sewage sludge, municipal wastewater or
farm slurries can provide very useful sources of nutrients to biomass crops – with
consequent positive growth responses. Perttu (1999), for example, has produced
figures to illustrate how municipal wastewater in Sweden provides macronutrients
in a ratio that almost exactly matches the calculated requirements of SRC willow
crops, as well as supplying potentially useful quantities of water.
Although nutrients and water supplied in sewage, wastewater, farm manures,
animal slurries and dirty water will boost plant growth generally, the results will
not necessarily be beneficial to wildlife. There are four main factors to consider.
1.
Direct ‘mulching’ effects of solid manures on ground vegetation: Irregular
distribution of manures will lead to suppression of ground vegetation beneath
thicker patches. This is likely to be a particular problem where application
rates are high. However, it could be considered less of a problem if the ground
flora is already very sparse, because of efficient weed control, or where flora
consists primarily of agricultural weed species of low conservation value.
2.
Effects of nutrient supply on ground flora species composition: The supply of
high levels of N and P in organic wastes will favour undesirable, highly
competitive weed species (e.g. nitrophilous species such as common nettle, U.
dioica) at the expense of more desirable, less competitive woodland species
(e.g. herb robert, Geranium robertianum; wood sorrel, Oxalis acetosella;
primrose, Primula vulgaris; common dog violet, Viola riviniana; bugle, Ajuga
reptans; and hedge woundwort, Stachys sylvatica). This factor can be
considered less significant if the alternative to organic waste application is the
supply of equivalent quantities of nutrients in inorganic fertilisers – as would
normally be the case.
3.
The indirect effects of increased nutrient supply to the crop: The application of
additional water and nutrients to any crop, if either was previously available at
below optimum levels, will produce a positive growth response – commonly
exhibited in higher shoot numbers, greater shoot lengths, and increased leaf
size and area. Taller, leafier energy crops will have high light interception
levels earlier in the season, and more effectively shade out plant species in the
ground and field layers.
4.
The introduction of new species to the planted area: There is a strong
possibility that the wastes being disposed of in biomass crops may include
plant seeds and fragments of stolons or rhizomes, and consequently be
responsible for the introduction of additional species to the flora. Sewage
sludge applications are commonly responsible for the introduction of tomato
plants, farmyard manure may contain seeds of grassland weeds (e.g. common
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BIOENERGY CROPS AND BIOREMEDIATION
chickweed and docks), and composted green household wastes may have
viable seeds of garden weed species.
Lynn et al. (1992) studied the short-term effects of applying different rates of
sewage sludge on the ground flora of a mixed-species woodland in Powys, Wales.
High rate applications (400 m3 ha-1, equivalent to a 4 cm deep mulch) in March
greatly reduced ground flora species diversity and vegetative cover, and delayed
plant emergence. The emergence of some species was delayed by as much as 10
weeks. Lower rate sludge applications (100 or 200 m3 ha-1) had less of an effect on
diversity, cover and emergence date. For this reason, Lynn et al. (1992) suggest
that lower rate applications might be more acceptable. Late season applications are
likely to have less of an impact on species diversity, as the sludge mulch will be
able to break down during the autumn and winter period.
Another, potentially negative effect of animal manure applications on biodiversity
(although positive in terms of crop production), is the possible inhibition of seed
germination by phenolic acids in the manures (Marambe & Ando, 1992).
Mastrota et al. (1989) reported that areas of a mixed oak forest in central
Pennsylvania that had been irrigated with chlorinated sewage effluent had
significantly lower densities of understorey trees and shrubs, but significantly
higher densities of herbs in the ground flora.
A Canadian study (Vasseur et al., 2000) of the effects of sewage sludge
applications to agricultural lands, on ground flora composition, found no
significant differences in species number between sludge-treated and untreated
sites. Vasseur et al. (2000) concluded that, although sewage sludge produced little
direct effect on ecological parameters, longer term effects on soil chemical factors
are likely to be reflected in plant community composition in the longer-term.
Soil biodiversity
The long-term effects of applying municipal wastewater on the spatial distribution
and biomass of soil micro-organisms were studied by Filip et al. (2000). Although
the soil samples were not taken from under a biomass crop, the results are still of
some relevance to this review.
One plot at the grassland study site, on a sandy soil in Germany, had been regularly
irrigated with wastewater for almost 100 years. Another plot had been irrigated
with wastewater, until about 20 years earlier. There was also a control plot that
had never received wastewater. During the sampling period, the irrigated plot
received approximately 2,000 mm wastewater (primary effluent) per annum. The
typical chemical composition of the wastewater was estimated to be 115 mg l-1
total C, 80 mg l-1 organic N, 71 mg l-1 NH4-N, 0.8 mg l-1 NO3-N, 14.2 mg l-1 P,
0.06 mg l-1 Cu, 1.1 mg l-1 Fe, 0.06 mg l-1 Mn, and 0.34 mg l-1 Zn. The pH of the
wastewater was around 8.0.
Results showed that numbers of bacteria, actinomycetes and fungi were generally
higher in the long-term irrigated soil. Organic particles in soil from this plot had
the highest microbial counts overall.
However, less nutrient-dependent
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BIOENERGY CROPS AND BIOREMEDIATION
oligotrophic bacteria were mostly associated with the silt + clay soil fraction,
irrespective of wastewater treatment. Slightly enhanced microbiological activity
was still detectable 20 years after the cessation of wastewater irrigation.
At a higher taxonomic level, earthworms grown in pots containing acid mine spoils
showed a positive growth response (increased surface area and mass) where the
spoil had been treated with sewage sludge (Pallant & Hilster, 1996).
Ground-dwelling invertebrates
Slater et al. (1997) found no effects of fertiliser or sludge treatment, or willow
cultivar, on ground-dwelling invertebrate populations, in terms of numbers of taxa
or individuals, within an SRC willow plantation in mid-Wales.
However, studies in other habitats have indicated that the application of organic
manures or sewage sludge can, in some instances, lead to increased invertebrate
populations. A field study by Larsen et al. (1996), in Ohio, USA, examined the
effects of 11 years of inorganic fertiliser or sewage sludge applications on the
ground beetle (Carabidae) fauna of an ‘old-field’ grassland community. Results
showed that plots treated with sewage sludge or inorganic fertiliser had
significantly more carabid species than untreated control plots. Another study
(Kielhorn et al., 1999), on a sandy mine spoil in Germany, showed that higher
numbers of ground beetle species and individuals were captured in plots
ameliorated with sewage sludge (compared with catches from plots treated with
compost, mineral fertiliser or untreated plots). All plots had been sown with a
grass cover (Secale multicaule) and planted with pine seedlings, and increases in
ground beetle populations were closely correlated with the amount of vegetative
cover.
Areas of a mixed oak forest in Pennsylvania that had been irrigated with
chlorinated sewage effluent had a greater biomass of invertebrates (gastropods and
annelids) than non-irrigated areas (Mastrota et al., 1989).
In some situations, livestock slurries may contain insecticides – a fact that is likely
to significantly affect their value as a substrate for invertebrates. A move away
from organophosphate-based sheep dips, towards synthetic pyrethroid-based
chemicals has been followed by new procedures for the disposal of these dips.
Pyrethroid sheep dips must now be diluted in animal slurry or water before
application to land. The negative environmental effects of slurries treated with
sheep dip may go beyond the direct impacts on pyrethroid-sensitive invertebrate
species. There may also be indirect effects on species higher in the food chain
through a) loss of invertebrate prey species and/or b) increased numbers of faecal
coliforms and pathogenic bacteria in slurries mixed with pyrethroid-based dips
(Semple et al., 2000).
Larsen et al. (1996) also showed that two of the most common carabid species,
Harpalus pensylvanicus and Poecilus lucublandus, accumulated Cd, Pb and Zn on
sewage-treated plots. However, concentrations recorded in these beetles were
lower than those present in the soil, indicating that they do not ‘bioconcentrate’
heavy metals. These are important findings, in relation to the potential
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BIOENERGY CROPS AND BIOREMEDIATION
bioaccumulation of heavy metals in the food chain and resultant risks that sewage
sludge applications pose to wildlife (see later discussion).
Canopy invertebrates
In contrast to their results for ground-dwelling invertebrates, Slater et al. (1997)
found that arboreal species were more abundant and taxa richness increased on
willow plots that were treated with P and K fertilisers and magnesian limestone,
when compared with a sewage sludge treatment. They suggest that this may be a
consequence of increased leaf areas and decreased foliar phenolic concentrations.
It has previously been shown that the phenolic compound salicylic acid is more
concentrated in leaves of willows grown in soils of low nutrient status, and that
high concentrations are toxic and a deterrent to leaf-feeding invertebrates (Pasteels
& Rowell-Rahier, 1992).
Of course, it does not necessarily follow that inorganic fertilisers will produce
larger leaf areas than organic wastes. The opposite situation may frequently apply
e.g. if slurries apply optimum (or higher) rates of N and other nutrients and useful
additional water.
Birds
No relevant information was found on the effects of waste applications to
bioenergy crops on bird populations. However, manures and other bulky organic
wastes will provide an important source of food for many invertebrates and thus,
indirectly, for some birds. Most bird species, even those that are herbivorous or
granivorous throughout most of their adult lives, have a predominantly invertebrate
diet prior to fledging. The availability of an abundant source of dung-feeding
insects, and their insect predators, during the spring and early summer nesting
period may improve the breeding success of birds in the vicinity of SRC or
Miscanthus plantations. Late winter or early spring manure applications to
recently harvested crops may provide a source of invertebrate food at the right
time, although the high density of most energy crops and their rapid rates of regrowth will probably make access very difficult for birds feeding second broods in
June or July.
Manures from cattle or sheep dosed with ‘worming’ insecticides such as
ivermectin may be toxic to invertebrates and, consequently, of little indirect benefit
to birds and insectivorous mammals. Wastes containing high concentrations of
heavy metals or organic toxins pose a potential risk to birds, particularly raptors,
through bioaccumulation of toxins in animal tissues.
Mammals
Lynn (cited by Slater et al., 1997), in an unpublished PhD study, found that sewage
sludge applications had no significant effects on populations of wood mice or bank
voles in broadleaved woodland. However, any resultant increase in invertebrate
populations as a result of manure, sludge or slurry applications is likely to be
beneficial to insectivorous mammals.
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BIOENERGY CROPS AND BIOREMEDIATION
A detailed study of small mammal populations within a mixed-oak forest in central
Pennsylvania (Mastrota et al., 1989) showed that wastewater (chlorinated sewage
effluent) had generally beneficial effects on small mammal populations. Small
mammal species richness was higher in areas of the forest that had been irrigated
with sewage effluent. Although Mastrota et al. recorded higher invertebrate
densities in sewage-treated areas, they suggest that the increased small mammal
diversity were more likely to be a consequence of the increased herb density in
these areas. This supposition appears to be supported by the findings that numbers
of white-footed mice (Peromyscus leucopus) and southern red-backed voles
(Clethrionomys gapperi) were significantly higher in sewage-treated areas,
whereas there were no significant differences between the numbers of two
insectivorous species (masked shrew, Sorex cinereus, and northern short-tailed
shrew, Blarina brevicauda) numbers in irrigated and non-irrigated zones.
Populations of white-footed mice also had significantly greater proportions of
adults on sites irrigated with sewage effluent.
The pollution of watercourses resulting from improper application of organic
wastes to energy crops is likely to have negative environmental effects overall,
although Vaughan et al. (1996) have shown how river eutrophication from sewage
pollution can have mixed effects on bat species foraging along that river. Their
results suggested that sewage discharges into the river had negative effects on
pipistrelles (Pipistrellus pipistrellus, 45 kHz phonic type), but positive effects on
Daubenton’s bats (Myotis daubentonii).
Another issue of considerable concern, however, is the possible accumulation of
heavy metals, and other toxins, in the internal organs of mammals when sewage
sludge or other wastes are repeatedly applied. This has been the subject of
research in a number of studies – although not in energy crops.
Sewage sludge can contain relatively high levels of heavy metals, such as Cd, Cr,
Cu, Pb, Ni and Zn – as well as organic pollutants and pesticides. Pig and poultry
manures have relatively high concentrations of Cu and Zn (Jongbloed & Lenis,
1998; Nicholson et al., 1999).
The toxicological effects of sewage sludge on young rats have been studied, under
laboratory conditions, by Bag et al. (1999). These include significant reductions in
levels of several enzymes, including liver alanine aminotransferase, liver succinate
dehydrogenase (SDH), serum lactate dehydrogenase (LDH) and muscle SDH; and
increases in liver and muscle LDH, serum and liver aspartate aminotransferase,
serum and muscle alkaline phosphatase, and brain and muscle acetylcholinesterase
activities.
The potentially toxic effects of heavy metal accumulation in the organs of
mammals inhabiting plantations regularly treated with sewage sludge or pig slurry
are likely to be greatest in predatory species – a point made by Bag et al. (1999)
and reinforced by various field studies. In one US study, Hegstrom & West (1989)
found higher levels of Cd, Cu, Pb and Zn in the livers and kidneys of insectivorous
Trowbridge’s shrews (Sorex trowbridgii) from forest plots treated with sewage
sludge. Despite high levels of heavy metals in shrews, no evidence was found of
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heavy metal-induced lesions in these tissues. Hegstrom and West also found that
sewage sludge applications increased levels of Cd and Pb in organs of
insectivorous shrew-moles (Neurotrichus gibbsii) and Cd in organs of granivorous
deer mice.
In another US study (Woodyard et al., 1987), a single application of municipal
sewage to forest stands in Michigan (10-year-old aspen, 50-year-old pine, 50-yearold mixed broadleaves and 70-year-old oak) produced no gross changes in
concentrations of Cd, Cr, Cu, Ni and Zn in the tissues of herbivorous small
mammals (or plants), when these were analysed 1-2 years later.
Nickelson & West (1996) examined the longer-term effects of sewage sludge
applications in forests on the Cd levels in kidneys of insectivorous shrews (Sorex
sp.) and omnivorous mice (Peromyscus sp.). Animals were captured at 15 sites in
western Washington, USA. Eight sites had been treated with anaerobically
digested biosolids (from mixed domestic and industrial sources), applied at varying
rates; 4, 11 or 15 years previously. The other seven, control, sites had not received
sewage applications. Results showed that Cd concentrations were higher in the
kidneys of shrews than mice, at all treatment sites. Except for sites receiving the
lowest rate applications, shrews from all sewage-treated forests had significantly
higher kidney Cd levels than shrews from controls. Cd concentrations in kidneys
of shrews trapped at the 11-year post-application sites were approximately
equivalent to those recorded from the same sites two years after sludge application.
These elevated levels of Cd in shrews did not, however, appear to be biologically
significant. The results for Peromyscus showed some significantly increased Cd
concentrations in mice from some sewage-treated sites, but this was not a
consistent trend, and Nickelson and West concluded that it was doubtful if sewage
biosolids had a significant long-term effect on these omnivores.
Campa et al. (1987) found that white-tailed deer (Odocoileus virginianus) and elk
(Cervus elaphus canadensis) in a clear-felled forest area browsed more heavily on
sludge-treated (9,980 kg ha-1) vegetation than in control areas. The vegetation in
sludge-treated areas had a higher crude protein content.
Deer grazing
predominantly in sludge-treated areas had higher heavy metal concentrations in
their tissues, but level were not considered to be high enough to threaten the health
of the deer or humans who might eat them.
There are numerous other publications relating to the accumulation of heavy
metals in mammals (e.g. Babish et al., 1982; Maly, 1984; Telford et al., 1984;
Anderson, 1985; Bray et al., 1985; Dressler et al., 1986; Alberici et al., 1989;
Brueske & Barrett, 1991).
The presence of pathogens (e.g Salmonella) in animal manures, improperly treated
sewage sludge (Wray & Callow, 1985) or slurries mixed with sheep dip (Semple et
al., 2000) may also pose a potential disease risk to grazing mammals and their
predators.
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BIOENERGY CROPS AND BIOREMEDIATION
2.4.3 Industrial wastes
There is a very wide range of waste materials from industrial processes that could
potentially be disposed of on bioenergy crops – or on sites intended subsequently
for bioenergy crop production. Those of greatest potential interest probably
include wastes from the paper industry, such as de-inked sludges.
Paper industry wastewater sludges are relatively low in heavy metals. Some fresh
sludges may be toxic (e.g. to bacteria and plants), but composting can readily
reduce toxicity making them suitable for application to agricultural land (Rantala et
al., 1999).
Vegetation
No relevant information was found on the effects of industrial waste applications to
bioenergy crops on ground vegetation. However, the direct effects of mulching on
the ground flora and indirect effects of nutrient enrichment on the crop and ground
flora described for agricultural manures will also be applicable for large volume
applications of industrial wastes.
Soil biodiversity
Erstfeld & Snow-Ashbrook (1999) studied the effects on soil-dwelling
invertebrates of low levels of polycyclic aromatic hydrocarbons (PAHs), in soils
from an abandoned industrial site. Their results suggested that higher PAH
concentrations (in the range 5.28 to 80.46 mg kg-1) were associated with increased
taxonomic diversity of nematodes, increased abundance of omnivorous and
predatory springtails (Collembola) and increased growth rates in earthworms
(Eisenia andrei). The same study showed a negative effect of increased PAH
concentration on the total abundance of mites (order Acarina).
It is not known what effects, if any, the much lower rates of PAHs found in animal
manures and slurries (see Section 2.3.2) might have on soil-inhabiting
invertebrates.
Krogh & Pedersen (1997) reported the results of a Danish study, which examined
the effects of applying dried and sterilised waste-water sludge, from a pesticide
factory, on the microarthropod fauna in the soil of a mature Norway spruce (Picea
abies) forest. The granulated sludge had a dry matter content of 93%, 8-10%
organic matter and contained 9.2% P and 25.5% CaCO3. The sludge granules also
contained small residues of phosphate triesters, pesticide residues, pyrimidines and
organic phosphoric acids. Results showed that the sludge significantly reduced
populations of springtails and mites. However, applications of inorganic fertiliser
had similar effects, although springtails were less affected by this treatment. Both
direct toxicity and changes in the microbial community as a result of increased
nutrient availability were suggested as probable causes of these population
declines. Overall, springtail populations were 30-35% lower in fertiliser and
sludge treatments than in control plots, one year after treatment application. One
springtail species, however, Isotoma notabilis, actually increased in numbers in
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BIOENERGY CROPS AND BIOREMEDIATION
response to sludge applications; to 2.6 times the control level (60% of total
springtail numbers) in the high dosage treatment (1.5 t ha-1).
Krogh & Pedersen had expected that the uneven distribution of sludge would result
in ‘protective sub-habitats’ - where soil fauna populations were unaffected – but
found no evidence of this.
Ground-dwelling invertebrates
No relevant information was found on the effects of industrial waste applications to
bioenergy crops on populations of ground-dwelling invertebrates. However, the
indirect effects on invertebrates resulting from changes in the ground flora crop,
due to mulching or nutrient enrichment – as described for agricultural manures will also be applicable for large volume applications of industrial wastes. Some
industrial wastes may contain chemicals that are directly toxic to invertebrates.
Canopy invertebrates
No relevant information was found on the effects of industrial waste applications to
bioenergy crops on populations of canopy invertebrates. However, wastes that are
nutrient-rich may increase the susceptibility of the energy crop to herbivorous
insects, although (as described for agricultural wastes and sewage sludge) this will
be largely irrelevant if the alternative is the application of similar quantities of
nutrients in inorganic fertilisers.
Birds
No relevant information was found on the effects of industrial waste applications to
bioenergy crops on bird populations. However, certain types of organic waste will
be attractive to birds, which either utilise the waste directly as a food source or feed
upon the invertebrate fauna foraging and breeding within the waste. For example,
Gabrey (1997) surveyed bird populations at different waste management facilities
in the USA and collected data that clearly demonstrates the attractiveness of
putrescible waste to gulls and other birds. In total, about 350 times more birds
were recorded at a putrescible landfill site than at five other sites, which included a
‘yard-waste compost facility and a ‘trash-transfer station.
Mammals
No relevant information was found on the effects of industrial waste applications to
bioenergy crops on mammal populations. However, as has been illustrated for
sewage sludge and animal manures, whenever wastes containing high levels of
heavy metals are applied there are risks of toxic materials accumulating in the
organs of animals that are resident within or utilise treated areas. Similar risks will
exist where energy crops are used to stabilise or bioremediate metal contaminated
sites, although the risks are likely to have been present prior to the establishment of
the energy crop. The dangers of toxicological damage are greatest to predatory
species, with the risks of ‘bioconcentration’, in the food chain. Relevant studies
include the work of Koeck et al. (1989), on heavy metal accumulation in small
mammals at a waste disposal site, and Andrews et al. (1984) on cadmium
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BIOENERGY CROPS AND BIOREMEDIATION
accumulation in the tissues of field voles and common shrews on grassland over
metalliferous mine waste.
Meyn et al. (1997) modeled the exposure risk of three representative bird and
mammal species, to assess the levels of hazard posed by 2,3,7,8tetrachlorodibenzo-p-dioxin (TCDD) in land applications of pulp and paper mill
sludge. Their results suggest that TCDD may, in some circumstances, pose
significant risks to terrestrial wildlife. Shrews were considered to be at most risk
because of their high rates of consumption of invertebrates that are expected to
accumulate TCDD from soils and vegetation at application sites.
Other research has looked at the potential toxicological effects in small mammal
species from exposure to hazardous materials, including petrochemical wastes
(McBee, 1991; Feuston & Mackerer, 1996), organochlorine insecticides (Rowley
et al., 1983) and even radioactive wastes (Arthur et al., 1986).
2.4.4 References
Alberici, T M; Sopper, W E; Storm, G L & Yahner, R H (1989). Trace metals
in soil, vegetation, and voles from mine land treated with sewage sludge.
Journal of Environmental Quality 18 (1). 115-120.
Anderson, D A (1985). Influence of sewage sludge fertilization on food habits of
deer in Western Washington. Journal of Wildlife Management 49 (1). 91-95.
Andrews, S M; Johnson, M S & Cooke, J A (1984). Cadmium in small
mammals from grassland established on metalliferous mine waste.
Environmental Pollution. Series A: Ecological and Biological 33 (2). 153162.
Arthur, W J; Markham, O D; Groves, C R; Keller, B L & Halford, D K
(1986). Radiation dose to small mammals inhabiting a solid radioactive waste
disposal area. Journal of Applied Ecology 23 (1). 13-26.
Babish, J G; Johnson, B; Brooks, B O; Lisk, D J (1982). Acute toxicity of
organic extracts of municipal sewage sludge in mice.
Bulletin of
Environmental Contamination and Toxicology 29 (4). 379-384.
Bag, S; Vora, T; Ghatak, R; Nilufer, I; D’Mello, D; Pereira, L; Pereira, J;
Cutinho, C & Rao, V (1999). A study of toxic effects of heavy metal
contaminants from sludge-supplemented diets on male Wistar rats.
Ecotoxicology and Environmental Safety 42 (2). 163-170.
Bray, B J; Dowdy, R H; Goodrich, R D & Pamp, D E (1985). Trace metal
accumulations in tissues of goats fed silage produced on sewage sludgeamended soil. Journal of Environmental Quality 14 (1). 114-118.
Brueske, C C & Barrett, G W (1991). Dietary heavy metal uptake by the least
shrew, Cryptotis parva. Bulletin of Environmental Contamination and
Toxicology 47 (6). 845-849.
Campa, H; Woodyard, D K & Haufler, J B (1987). Deer and elk use of forages
treated with municipal sewage sludge. In: The forest alternative for treatment
and utilization of municipal and industrial wastes (Eds. D W Cole, C L Henry
& W L Nutter). University of Washington Press, Seattle. 188-198.
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BIOENERGY CROPS AND BIOREMEDIATION
Coates, A & Say, A (1999). Ecological assessment of short rotation coppice.
ETSU B/W5/00216/REP/1. Contract report by Environmental Resources
Management for the Department of Trade and Industry.
Dressler, R L; Storm, G L; Tzilkowski, W M & Sopper, W E (1986). Heavy
metals in cottontail rabbits on mined lands treated with sewage sludge.
Journal of Environmental Quality 15 (3). 278-281.
Erstfeld, K M & Snow-Ashbrook, J (1999). Effects of chronic low-level PAH
contamination on soil invertebrate communities. Chemosphere 39 (12). 21172139.
Filip, Z; Kanazawa, S & Berthelin, J (2000). Distribution of microorganisms,
biomass ATP, and enzyme activities in organic and mineral particles of a longterm wastewater irrigated soil. J. Plant Nutr. Soil Sci. 163. 143-150.
Feuston, M H & Mackerer, C R (1996). Developmental toxicity study in rats
exposed dermally to clarified slurry oil for a limited period of gestation.
Journal of Toxicology and Environmental Health 49 (2). 207-220.
Fuller, R J & Green, G H (1998). Effects of woodland structure on breeding bird
populations in stands of coppiced lime (Tilia cordata) in western England over
a 10-year period. Forestry 71(3). 199-218.
Gabrey, S W (1997). Bird and small mammal abundance at four types of wastemanagement facilities in northeast Ohio. Landscape and Urban Planning 37
(3-4). 225-235.
Hegstrom, J & West, S D (1989). Heavy metal accumulation in small mammals
following sewage sludge application to forests. Journal of Environmental
Quality 18 (3). 345-349.
Jongbloed, A W & Lenis, N P (1998). Environmental concerns about animal
manure. In: Proceedings of ‘Nutrient Management Procedures to Enhance
Environmental Conditions’ symposium. July 1997; Nashville, USA. Journal
of Animal Science 76 (10). 2641-2648.
Kielhorn, K –H; Keplin, B & Hüttl, R F (1999). Ground beetle communities on
reclaimed mine spoil: Effects of organic matter application and revegetation.
In: Proceedings of the international symposium Organic matter application
and turnover in disturbed terrestrial ecosystems (Ed. D Vetterlein). Cottbus,
Germany. November 1997. Plant and Soil 213 (1-2). 117-125.
Koeck, M; Schaffler, R; Sixl, W; Pichler-Semmelrock, F P; Kosmus, W &
Marth, E (1989). Accumulation of heavy metals in animals. III: Heavy metal
accumulation in small mammals at a waste disposal site. Journal of Hygiene,
Epidemiology, Microbiology and Immunology 33 (4). 536-541.
Krogh, P H & Pedersen, M B (1997). Ecological effects assessment of industrial
sludge for microarthropods and decomposition in a spruce plantation.
Ecotoxicology and Environmental Safety 36. 162-168.
Larsen, K J; Purrington, F F; Brewer, S R & Taylor, D H (1996). Influence of
sewage sludge and fertilizer on the ground beetle (Coleoptera: Carabidae)
fauna of an old-field community. Environmental Entomology 25 (2). 452459.
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Lynn, S F; Slater, F M & Randerson, P F (1992). The ecological impact of
sewage sludge applications on woodland vegetation. Aspects of Applied
Biology 29, Vegetation management in forestry, amenity and conservation
areas. 383-388.
Maly, M S (1984). Survivorship of meadow voles, Microtus pennsylvanicus, from
sewage sludge-treated fields. Bulletin of Environmental Contamination and
Toxicology 32 (6). 724-731.
Marambe, B & Ando, T (1992). Phenolic acids as potential seed germinationinhibitors in animal waste composts. Soil Science and Plant Nutrition (Tokyo)
38 (4). 727-733.
Mastrota, F N; Yahner, R H & Storm, G L (1989). Small mammal communities
in a mixed oak forest irrigated with wastewater. Am. Midl. Nat. 122 (2). 388393.
McBee, K (1991).
Chromosomal aberrations in native small mammals
(Peromyscus leucopus) at a petrochemical waste disposal site. II: Cryptic and
inherited aberrations detected by a G-band analysis.
Environmental
Toxicology and Chemistry 10 (10). 1321-1329.
Meyn, O; Zeeman, M; Wise, M & Keane, S E (1997). Terrestrial wildlife risk
assessment for TCDD in land-applied pulp and paper mill sludge: wildlife
ecotoxicology. In: Proceedings of symposium on Wildlife Ecotoxicology.
Copenhagen, Denmark, June 1995. Environmental Toxicology and Chemistry
16 (9). 1789-1801.
Nicholson, F A; Chambers, B J; Williams, J W & Unwin, R J (1999). Heavy
metal contents of livestock feeds and animal manures in England and Wales.
Bioresource Technology 70. 23-31.
Nickelson, S A & West, S D (1996). Renal cadmium concentrations in mice and
shrews collected from forest lands treated with biosolids. Journal of
Environmental Quality 25 (1). 86-91.
Pallant, E & Hilster, L M (1996). Earthworm response to 10 weeks of incubation
in a pot with acid mine spoil, sewage sludge and lime. Biology and Fertility of
Soils 22 (4). 355-358.
Pasteels, J M & Rowell-Rahier, M (1992). The chemical ecology of herbivory
on willows. In: Proceedings of Botanical Society of Edinburgh symposium on
Willow; Edinburgh, UK; 27-29 September 1991. Proc. Royal Society of
Edinburgh (Sect B) 98 (0). 63-73..
Perttu, K (1999). Environmental and hygienic aspects of willow coppice in
Sweden. Biomass and Bioenergy 16. 291-297.
Rantala, P-R; Vaajasaari, K; Juvonen, R; Schultz, E; Joutti, A & MäkeläKurtto, R (1999). Composting of forest industry wastewater sludges for
agricultural use. In: Forest Industry Wastewaters VI (Ed. A Luonsi). Selected
proceedings of the 6th IAWQ Symposium on Forest Industry Wastewaters,
Tampere, Finland; 6-10 June 1999. Wat. Sci. Tech. 40 (11-12). 187-194.
Reddersen, J (2001). SRC-willow (Salix viminalis) as a resource for flowervisiting insects. Biomass and Bioenergy 20. 171-179.
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BIOENERGY CROPS AND BIOREMEDIATION
Rowley, M H; Christian, J J; Basu, D K; Pawlikowski, M A & Paigen, B
(1983). Use of small mammals (voles) to assess a hazardous waste site at
Love Canal, Niagara Falls, New York.
Archives of Environmental
Contamination and Toxicology 12 (4). 383-397.
Sage, R B & Robertson, P A (1996). Factors affecting songbird communities
using new short rotation coppice habitats in spring. Bird Study 43. 201-213.
Sage, R B & Tucker, K (1997). Invertebrates in the canopy of willow and poplar
short rotation coppices. Aspects of Applied Biology 49, Biomass and energy
crops. 105-111.
Sage, R B & Tucker, K (1998a). The distribution of Phratora vulgatissima
(Coleoptera: Chrysomelidae) on cultivated willows in Britain and Ireland.
Eur. J. For. Path. 28. 289-296.
Sage, R & Tucker, K (1998b). Integrated crop management of SRC plantations
to maximise crop value, wildlife benefits and other added value opportunities.
ETSU/B/W2/00400/REP.
Semple, K T; Hughes, P; Langdon, C J & Jones, K (2000). Impact of synthetic
pyrethroid sheep dip on the indigenous microflora of animal slurries. FEMS
Microbiology Letters 190 (2). 255-260.
Šlapokas, T & Granhall, U (1991). Decomposition of willow-leaf litter in a
short-rotation forest in relation to fungal colonization and palatability for
earthworms. Biology and Fertility of Soils 10. 241-248.
Slater, F M; Hodson, R W; Randerson, P F & Lynn, S F (1997). Some
environmental impacts of short rotation willow coppice. In: Making a
Business from Biomass in Energy, Environment, Chemistry, Fibres and
Materials (Eds R P Overend & E Chornet). Proceedings 3rd Biomass
Conference of the Americas, Montreal, Quebec, Canada, August 24-29, 1997.
Pergamon, Elsevier Science. 29-37.
Telford, J N; Babish, J G; Johnson, B E; Thonney, M L; Currie, W B; Bache,
C A; Gutenmann, W H & Lisk, D J (1984). Toxicological studies with
pregnant goats fed grass-legume silage grown on municipal sludge-amended
topsoil. Archives of Environmental Contamination and Toxicology 13 (5).
635-640.
Vasseur, L; Cloutier, C & Ansseau, C (2000). Effects of repeated sewage sludge
application on plant community diversity and structure under agricultural field
conditions on Podzolic soils in eastern Quebec. Agriculture, Ecosystems &
Environment 81 (3). 209-216.
Vaughan, N; Jones, G & Harris, S (1996). Effects of sewage effluent on the
activity of bats (Chiroptera: Vespertilionidae) foraging along rivers.
Biological Conservation 78 (3). 337-343.
Vaughan, N; Jones, G & Harris, S (1997). Habitat use by bats (Chiroptera)
assessed by means of a broad-band acoustic method. Journal of Applied
Ecology 34. 716-730.
Woodyard, D K; Campa, H & Haufler, J B (1987). The influence of forest
application of sewage sludge on the concentration of metals in vegetation and
small mammals. In: The forest alternative for treatment and utilization of
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municipal and industrial wastes (Eds. D W Cole, C L Henry & W L Nutter).
University of Washington Press, Seattle. 199-205.
Wray, C & Callow, R J (1985). A note on potential hazards to animals grazing
on pasture improperly treated with sewage sludge. Journal of Applied
Bacteriology 58 (3). 257-258.
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CHAPTER 3
BIOREMEDIATION OF CONTAMINATED SITES
PETER NIXON
3.1
INTRODUCTION
The public welfare concern over the effects of environmental pollution has
increased substantially since the Industrial Revolution, mainly as a consequence of
an enhanced understanding of the risk to human health. Much concern has
concentrated on the visible effects of pollution, but the hidden effects are also of
great importance. Large concentrations of inorganic pollutants (i.e. trace elements)
in soils can be of either natural or anthropogenic origin. Natural sources are the
result of weathering processes, volcanic activity and natural fires in which plants or
wood burn. The main sources of pollution are from the burning of fossil fuels,
mining and smelting of metalliferous ores, inorganic and organic municipal waste
and sewage, and residues from materials used in agriculture.
The necessity to decontaminate polluted sites is recognised, both socially and
politically, because of the increasing importance placed on environmental
protection and human health. As the number of sites and levels of contamination
rise, so does the need to develop effective and affordable methods for
decontamination (Lombi et al., 1998).
3.2
PHYTOREMEDIATION
The use of plants for soil reclamation has begun to be looked at as an alternative to
physical or chemical processes. The term ‘phytoremediation’ is used to describe
techniques in which plants are used for the in situ treatment of soils polluted by
chemicals or radioactivity. There are four fundamental processes by which plants
can be used to remediate soils contaminated with trace elements (Salt et al., 1996):
1.
2.
3.
4.
Phyto-immobilisation – where plants are used to prevent the movement and
transportation of dissolved contaminants within the soil.
Phyto-stabilisation - where pollutant-tolerant plants are used to mechanically
stabilise polluted soils, and to prevent bulk erosion and airborne transport to
other environments. In addition, both phyto-immobilisation and phytostabilisation may reduce contaminated run-off, due to higher evaporation
rates relative to bare soil.
Phyto-extraction – where plants are used to extract metallic and organic
compounds from the soil into plant tissue.
Phyto-volatilisation – where specialised enzymes transform and volatilise
contaminants in a plant-microbe-soil system.
Phytoremediation is particularly suited to large sites that have relatively low levels
of contamination, but still sufficiently contaminated to restrict their use and limit
development potential. This low-cost option can also be used to stabilise the site
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BIOENERGY CROPS AND BIOREMEDIATION
until funds and/or techniques have become available to remediate the contaminated
land.
The production of biomass fuel crops on degraded or contaminated land has
several advantages as a means of site remediation.

Biomass crop production will not only bring such sites into economic use,
but may also improve their aesthetic value, and potentially lead to long-term
remediation, through heavy metal removal in the harvested crop (D. RiddellBlack - pers. com.).

Some current biomass crops, such as Salix, Populus and Miscanthus are
pioneer species and are adapted to harsh growing conditions - which typify
derelict sites. For example, Miscanthus invades derelict, abandoned and
volcanic areas in its native Japan (personal observation).

Establishment and management costs are low compared with chemical
washing.

The contaminates are contained on site, rather than being removed and
transported to another site.

The energy conversion offers a method for concentrating the metals
contained within the biomass.

Lastly the production of a low risk, non-food crop on otherwise unproductive
land can play a role in the rejuvenation of the local economy (D. RiddellBlack - pers. com.).
In Britain, Punshon et al. (1995) found that species of willows accumulated
cadmium in the leaves and stems. In Sweden, again with willows, it was found
that the average stem contained 2.1 mg dm kg-1, and it was estimated that 21 g ha-1
annually could be removed from the soil, with an annual crop yield of 10 t DM
ha-1. This corresponded to the removal of 3-4% of the plant available cadmium
(Ostman, 1994).
Under controlled conditions in the laboratory it was observed that Betula pendula
was able to accumulate high concentrations of cadmium (Goransson, 1994). It was
estimated that up to 1.5 kg Cd ha-1 yr-1 could be removed from the soil with no
detrimental effect on the growth of the plant. The vast majority (95%) of this Cd
was found to be in the roots. This study estimated that there was also a significant
annual uptake of other metals, such as Al (25 kg ha-1 yr-1), Cu (1.5 kg ha-1 yr-1), Fe
(5 kg ha-1 yr-1), and Mn (50 kg ha-1 yr-1). Again in Sweden, Ericson (1994) found
that Salix spp. are capable of accumulating 1.3 g Cd ha-1 annually.
Punshon et al. (1995) also investigated the resistance to copper toxicity in various
British willows and found that there were significant differences in root length, the
number of lateral roots and the pattern of metal uptake between species, hybrids
and clones. This study also found that Salix caprea, S. cinerea and their hybrids,
and S. viminalis appear to grow best in elevated copper solution. Other studies
have also demonstrated the ability of Salix species to tolerate polluted and hostile
environments. For example, S. caprea will grow on lead mine spoils (Eltrop et al.,
1991) and the hybrids of S. caprea, S. cinerea and S. viminalis will grow on metal
contaminated river silt (Mang & Reher, 1992).
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BIOENERGY CROPS AND BIOREMEDIATION
The uptake of heavy metals by twenty different varieties of willow showed that
there were significant differences in the concentrations of Pb, Cr, Cu and Ni in the
above ground tissues (Riddell-Black, 1997). However, the varieties tested
appeared to fall into two distinct groups:

Those that accumulated Cu or Ni, in the above ground biomass, and as a
consequence suffered substantially reduced yields.

Those that did not accumulate Cu or Ni, and so suffered no yield reduction.
The willow varieties tested seemed able to tolerate high concentrations of Zn or Cd
in bark tissue, without yields being influenced. From this study, there appears to
be two scenarios for the use of willow SRC on contaminated land:
1.
Planting of non-accumulating varieties (e.g. Salix ‘Rosewarne White’ and S.
spaethii), to stabilise the metal losses via wind and water erosion. This option
allows only limited heavy metal removal, but does permit the use of the wood
for fuel.
2.
Planting a variety such as S. burjatica ‘Germany’, which appears to remove
Cu, Zn, Cd and Ni from the soil in significant quantities, having both high
metal concentrations in the above ground tissue and high yields. However, it
may not be possible to use this wood for fuel with the current technology.
Carol et al. (1996) used grey alder (Alnus incana), silver birch (Betula pendula)
and Scots pine (Pinus sylvestris) in the phytoremediation of an area of steelworks’
slag, in Lanarkshire - contaminated with copper, nickel and zinc. They found that
these three tree species could all be successfully grown on the moderately alkaline
slag. Trees grew best on slag that had been treated with sewage cake prior to
planting. Grey alder had the best growth rates. However, the trees only managed
to accumulate less than 1% of the top soil metal contamination in the above ground
biomass. This could have been due to the high pH and the chemical form of the
metals in the waste. The scheme was, nonetheless, judged a success as it greened
the site, created a wildlife habitat, improved soil fertility and stabilised the metal
contaminants.
Scott et al.. (1995) looked at the uptake of copper, chromium, lead and zinc in
birch, willow and sycamore (Acer pseudoplatanus) growing on two sites; one an
ex-chromium works and the other an ex-ammunition factory. These trees were not
able to significantly accumulate the metals under investigation, although zinc was
the most mobile and was detected at reasonably high concentrations in the above
ground biomass, the concentrations being highest in the bark and leaves. Changes
in zinc concentration in different parts of the tree took place over the growing
season, with the suggested optimum time for harvest being the winter period. The
other metals were only found at high concentrations in the tree roots; a positive
correlation between zinc uptake and zinc concentration in the soil was found.
Steer & Baker (1997) investigated the use of three trees - Populus deltoides x
trichocarpa (P. x interamericana) ‘Beaupré’, common alder (Alnus glutinosa) and
grey willow (Salix cinerea) – grown as SRC biomass crops, to re-generate two old
reclaimed coal tips in Wales. All three species showed very high rates of survival.
On the better of the two sites, poplar produced mean yields of up to 6 t DM ha-1 yr1
(excluding dead trees), at the first harvest, in plots pre-treated with an application
of 300 mm raw sewage sludge (which was applied below the top 1 m of colliery
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BIOENERGY CROPS AND BIOREMEDIATION
spoil). The authors suggest that their results indicate the potential for SRC poplar
grown on old coal spoil to produce yields comparable to those from some lowland
agricultural sites, but only under the right conditions i.e. with ready availability of
primary nutrients and a lack of compaction. The other two species (and poplar in
other treatments) were less impressive in terms of yield, but alder was capable of
becoming established and growing with little if any pre-treatment.
This work (Steer & Baker, 1997) highlighted the criteria for successful
establishment of trees on this type of land:

choice of the correct species or clones for local site conditions

thorough site preparation and, especially, avoidance of compaction

availability of nutrients

good site management, especially the use of effective weed control and
fencing.
Steer & Baker (1997) concluded that the use of SRC on re-claimed coal sites had
real potential, both for the production of biomass, and also for the benefits of the
environment.
A pot study by Wilkins (1997) on the uptake of copper, arsenic and zinc by
Miscanthus sp. grown on metalliferous soils, showed little difference in uptake on
polluted or unpolluted soils. This study did demonstrate that Miscanthus was able
to survive and to grow on highly polluted soils, and that the use of inorganic
fertiliser and/or lime greatly improved the yield of Miscanthus with very little
influence on the uptake of metals. The work concluded that there was little cause
for concern in the growing of Miscanthus as a biofuel on soil polluted by mining.
Another pot study followed the uptake and mobilisation of cadmium, zinc and
copper in Miscanthus x giganteus. It showed that levels of Cu and Zn were mainly
confined to the roots and rhizomes, but concentrations in plants did not show any
increase in uptake with increased soil supply. This is in complete contrast to
cadmium concentrations, which did increase with increases in soil supply. Levels
of cadmium were found to be highest in the roots and rhizomes (S. Wilson - pers.
comm.).
The use of Miscanthus in Portugal as an accumulator of heavy metals was studied,
using pots. It was found that the level of heavy metal concentration in the soil
negatively affected growth and productivity. The results indicated that Miscanthus
was able to accumulate and remove heavy metals (Cd, Cr, Cu, Ni, Pb & Zn) from
the soil in the below-ground fraction of the plant, but that the aerial fraction did not
significantly accumulate the heavy metals (Fernando et al., 1996). The work was
repeated in the field, comparing fields with different levels of contaminants.
Results showed that productivity, plant height and stem numbers all increased with
increased levels of sewage sludge (up to 100 t ha-1), but levels of heavy metals (Cd,
Cr, Cu, Ni, Pb & Zn) in the above-ground biomass did not significantly increase
with increased metal concentrations in the soil. The ash, nitrogen and phosphorus
content of the above ground biomass did increase with increased levels of
contaminants. However, the below-ground material did accumulate higher
concentrations of heavy metals with increased levels of contaminants (A Fernando
- pers. com.).
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BIOENERGY CROPS AND BIOREMEDIATION
The suitability of perennial grasses from the botanical garden of the Plant Breeding
and Acclimatization Institute (PBAI), Bydgoszcz, Poland was assessed, in order to
see if they could be used to phyotremediate a coal dump in Poland (Majtkowski &
Majtkowska, 1997). The use of grasses are favoured in Poland as there is concern
that deep rooting trees may cause oxidation of pyrite, resulting in spontaneous
combustion. Two of the most promising species were found to be Miscanthus
sacchariflorus and Spartina michauxiana. It was also found that the incorporation
of biogel (water absorbent product) into the soil increased the absorption ability of
plants in poor soils and improved the plants' ability to utilise moisture. The study
concluded that the use of grass offered one of the most effective methods of
remediating the coal-mine spoils of the Katowice region.
AEA Technology, in collaboration with Border Biofuels and Derbyshire County
Council, have investigated the use of SRC as a low cost route to the remediation of
coal spoil sites - and identified a number of benefits. These include the rapid
capture and greening of a site, site stabilisation due to the extensive network of
fibrous roots, high productivity, leaf fall leading to increased organic matter and
the creation of soil, and the control of contaminated run-off. The use of trees may
benefit the local area in other ways such as improving the visual amenity of the
area and the creation of new woodlands with increased public access and
recreational activities. Their methods are based on the incorporation of organic
material (e.g. sewage and other organic sludges) into the surface of the
contaminated soil, prior to the planting of SRC (F Dumbleton - pers. com.).
The recently completed, EC-funded, BIORENEW project examined the potential
for the use of biomass fuel crops in the bioremediation and economic renewal of
industrially degraded land. The main objective of the BIORENEW project was to
develop a system for the rehabilitation of heavy metal contaminated land using
biomass fuel crops, which brings a net environmental benefit and contributes to the
economic regeneration of areas suffering industrial decline. A number of field
screening trials of Salix, Phalaris and Eucalyptus in the UK, Sweden and Spain
were undertaken; along with the screening of 150 Salix clones, 20 Phalaris clones
and 10 Miscanthus for the uptake of heavy metals.
The use of vegetation on completed landfill sites has been studied for a number of
years. Ettala (1988), in southern Finland, concluded that SRC could be
successfully established on sanitary landfills. Of the five Salix spp. planted, the
most productive was S. aquatica. The SRC on the landfill was managed in the
same way as a conventional SRC crop on an arable soil. The study also found that
the best results were achieved when the substrate had a high humus content and a
thickness of at least 0.2-0.3 m. The planting of SRC on landfill notably improved
the landscape of the site but also increased evapotranspiration.
The use of vegetation on landfill sites restoration is also important, in that it
provides shelter and reinforcement for the capping system and contributes to its
basic functioning. However, it may cause a potential risk, due to the damage that
some tree species roots may cause by penetrating some types of sealers (McDonald
et al., 1997). Work in Hong Kong on the restoration of completed landfill sites
(Chan, 1997), using Acacia confusa and Casuarina equisetifolia, found that these
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BIOENERGY CROPS AND BIOREMEDIATION
two N2-fixing trees had a rooting system which was confined to the upper 15 cm of
the topsoil and did not penetrate any deeper; thus the risk of penetrating the 1 m
cap was very low. It was also possible to establish these species without the need
for imported planting medium, which could increase the cost of remediation.
3.3
POTENTIAL ENVIRONMENTAL PROBLEMS
The two main areas of concern with regard to planting a biomass crop for the
phytoremediation of contaminated land are the possible contamination of the wider
environment via the food chain and of the release of metals following combustion.
3.3.1 Bioaccumulation of pollutants
Studies on the transfer of heavy metals by animals feeding on contaminated sites
and the subsequent toxic hazards to birds and other top predators is incomplete, but
the main route is likely to be via soil feeding animals (Punshon & Dickinson,
1997). The ability of earthworms to accumulate cadmium is a good example, and
thus earthworm-eating birds may be susceptible. However, most birds feed over a
large area and on a range of invertebrates, and unless they feed exclusively on a
contaminated site their chances of suffering from toxic effects are unlikely
(Furness, 1996). Small mammals may also accumulate high levels of cadmium
and lead, so top predators feeding exclusively on small mammals from
contaminated sites may be exposed to elevated levels. These issues are discussed
more fully in Section 3.5 of this review.
3.3.2 Release of metals during combustion
Besides the possible threat to the environment from the combustion of biomass,
there are also additional operating costs due to the higher specification of the flue
gas cleaner to ensure satisfactory metal removal. It has been estimated that the cost
of flue gas cleaning is between 5 and 15% of the capital cost of a combined heat
and power (CHP) system, but less than 1% of the annual running costs when using
conventional fuel sources. Therefore, any increase in the level of metals in
biomass fuel material may have an implication in the capital and annual costs for
systems using such fuels.
There is a further problem, as the resulting ash may have to be disposed of in a
landfill site at a cost of between £5 and £15 per tonne, rather than being used as a
good source of P and K for biomass and other crops. If the ash was designated as
hazardous or as special waste this would further increase the disposal cost to
between £50 and £150 per tonne (Riddell-Black et al., 1996).
Narodoslawsky et al. (1996) have suggested some solutions to this problem.
Firstly the mixing of fly-ash and bottom in equal quantities to produce a blended
ash which may be recycled on to the land. Secondly (the long-term solution) the
use of different temperatures in a biomass combustion plant to concentrate volatile
metals, such as cadmium, in a very small portion of the ash. This small portion of
ash could then be disposed of safely in a landfill site, and the remaining clean ash
used in agriculture.
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BIOENERGY CROPS AND BIOREMEDIATION
3.4
REFERENCES
Carol, S A; Hipkin, A J & Davidson, B (1996). Phytoremediation - a feasible
option at Lanarkshire steelworks? In: Heavy Metals and Trees (Ed. I
Glimmerveen). Institute of Chartered Foresters, Edinburgh. 51-62.
Chan, Y. S. G (1997). Root growth patterns of two nitrogen-fixing trees under
landfill conditions. Land Contamination & Reclamation 5. 55-62.
Eltrop, L; Brown, G; Hinchee, R E & Olfenbuttel, R F (1991). Lead tolerance
of Betula and Salix in mining areas of Mechernich, Germany. Plant Soil 131.
275-285.
Ericson, S-O (1994). Salix can remove cadmium from arable land - technical and
infra-structural implications. Proceedings of a study tour, conference and
workshop in Sweden, June 1994. Eds. P Aronsson & K Perttu. Swedish
University of Agricultural Sciences Report 50. 173-174.
Ettala, M O (1988). Short-rotation tree plantations at sanitary landfills. Waste
Management & Research 6. 291-302.
Fernando, A; Duarte, P & Oliveira, J F S (1996). Bioremoval of heavy metals
from soil by Miscanthus sinensis giganteus. Biomass for Energy and the
Environment. Proceedings of the 9th European Bioenergy conference. 531536.
Furness, R W (1996). Transfers of heavy metals from contaminated land to top
predators; implications for birds for growing trees for soil amelioration. In:
Heavy Metals and Trees (Ed. I Glimmerveen). Institute of Chartered
Foresters, Edinburgh. 107-122.
Goransson, A & Philippot, S (1994). The use of fast growing trees as ‘metalcollectors’. Proceedings of a study tour, conference and workshop in Sweden,
June 1994. Eds. P Aronsson & K Perttu. Swedish University of Agricultural
Sciences Report 50. 129-132.
Lombi, E; Wenzel, W W & Adriano, D C (1998). Soil contamination, risk
reduction and remediation. Land Contamination & Reclamation 6. 183-197.
Majtkowski, W & Majtkowska, G (1997). Suitability of ornamental perennial
grasses to reclamation of coal-mining dump. Ecological aspects of breeding
fodder crops and amenity grasses. Plant Breeding and Acclimitization
Institute, Bydgoszcz, Poland. 249-252.
Mang, F W C; Reher, R (1992). Heavy metal resistance clones of willows from
polluted areas useful for land restoration programmes. Proceedings of the
Royal Society, Edinburgh 98B. 244.
McDonald, C; Meggyes, T & Simmons, E (1997). Landfill capping: engineering
and restoration, 1. Structure and function of landfill capping system. Land
Contamination & Reclamation 5. 89-97.
Narodoslawsky, M & Obernberger, I. (1996). From waste to raw material - the
route from biomass to wood ash for cadmium and other heavy metals. Journal
of Hazardous Materials 50. 157-168.
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BIOENERGY CROPS AND BIOREMEDIATION
Ostman, G (1994). Cadmium in Salix - a study of the capacity of Salix to remove
cadmium from arable soils. Proceedings of a study tour, conference and
workshop in Sweden, June 1994. Eds. P Aronsson & K. Perttu. Swedish
University of Agricultural Sciences Report 50. 153-155.
Punshon T; Lepp, N W & Dickinson, N M (1995). Resistance to copper toxicity
in some British willows. Journal. of Geochemical Exploration 52, 259-266.
Punshon, T & Dickinson, N M (1997). Mobilisation of heavy metals using shortrotation coppice. Aspects of Applied Biology 49. Biomass and Energy Crops.
285-292.
Punshon, T; Dickinson, N M & Lepp, N W (1996). The potential of Salix clones
for bioremediating metal polluted soil. In: Heavy Metals and Trees (Ed. I
Glimmerveen). Institute of Chartered Foresters, Edinburgh. 93-104.
Riddell-Black, D M; Pulford, I D & Stewart, C (1997). Clonal variation in
heavy metal uptake by willow. Aspects of Applied Biology 49. Biomass and
Energy Crops. 327-334.
Riddell-Black, D M; Rowlands, C & Snelson, A (1996). The take up of heavy
metals by wood fuel crops - implications for emission and economics.
Biomass for Energy and the Environment. Proceedings of the 9th European
Bioenergy conference 3. 1754-1759.
Salt, D E; Blaylock, M; Kumar, N P B A; Dushenkov, V; Ensley, B D; Chet, L
& Raskin, L (1996). Phytoremediation: a novel strategy for the removal of
toxic metals from the environment using plants. Biotechnology 13. 468-474.
Scott, McG D; Duncan, H J; Pulford, I D & Wheeler, C T (1996). Uptake of
heavy metals from contaminated soil by trees. In: Heavy Metals and Trees
(Ed. I Glimmerveen). Institute of Chartered Foresters, Edinburgh. 171-176.
Steer, P & Baker, R M (1997). Colliery spoil, sewage and biomass - potential for
renewable energy from wastes. Aspects of Applied Biology 49. Biomass and
Energy Crops. 300-305.
Wilkins, C (1997). The uptake of copper, arsenic and zinc by Miscanthus environmental implications for use as an energy crop. Aspects of Applied
Biology 49. Biomass and Energy Crops. 335-340.
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CHAPTER 4
WASTE UTILISATION AND BIOREMEDIATION: THE ROLE OF
GENETICALLY MODIFIED ORGANISMS
NATASHA SMITH
4.1
INTRODUCTION
The genetic modification of biomass crops to facilitate increased pollutant uptake,
transport, accumulation and tolerance offers the potential to dramatically increase
the effectiveness of phytoremediation of organic and metal pollutants from
contaminated sites. Optimisation of performance using traditional plant breeding
techniques is limited to the genetic diversity found within the species itself. Recent
advances in molecular biology, in particular genetic engineering, make it possible,
in theory, to introduce virtually any gene into any organism. In order to achieve
such a goal, it is necessary to understand the mechanisms employed by organisms
to prevent poisoning, as a result of exposure to high levels of metals and organic
compounds. This chapter will examine the current understanding of these
mechanisms, their application to phytoremediation and the progress of
transformation technology with respect to biomass crops.
The differing nature of inorganic and organic pollutants requires different
mechanisms to facilitate remediation. Heavy metals and radionuclides are classed
as inorganic pollutants and examples include cadmium, cesium, chromium, lead,
mercury, strontium and uranium. These pollutants are immutable in terms of
biological or physical processes, excluding nuclear fission and fusion. Effective
phytoremediation of such substances are therefore, limited to
(i)
the extraction, translocation and sequestration of the toxic ions to plant
tissues; facilitating the prevention of leaching from the site, or removal
from the site by harvesting the plant material, or
(ii)
conversion of the element to a less toxic chemical species, for example:
transformation of chromium VI to chromium III (James, 1996).
In contrast, many organic pollutants can be reduced to relatively non-toxic
compounds and even CO2 and water. Targets for phytoremediation include
polychlorinated biphenyls (PCBs), polycyclic aromatic hydrocarbons (PAHs),
trichloroethylene (TCE) and trinitrotoluene (TNT).
Plants have many endogenous genetic, biochemical and physiological systems that
make them ideal for the remediation of soils and water. Pollutants are often taken
up by the same mechanisms that adsorb, transport and translocate nutrients. Once
adsorbed and transported, pollutants can either hyperaccumulate or undergo some
form of transformation.
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BIOENERGY CROPS AND BIOREMEDIATION
4.2
PHYTOREMEDIATION OF INORGANIC POLLUTANTS
4.2.1 Root uptake
Plant root systems offer a large surface area with which to adsorb nutrients from
the surrounding soil matrix (Boyajian & Carreira, 1997). Inorganic pollutants also
become bound to the root surface during adsorption. In liquid media, the roots of
Brassica juncia have been shown to contain 500 times more Cd(II), Ni(II), Pb(II)
and Sr(II) than the surrounding liquid (Salt et al., 1995; Salt & Kramer, 1999).
Similarly, sunflower roots can accumulate uranium such that it is 30,000 times
more concentrated than the surrounding liquid media (Dushenkof et al., 1997).
These levels of accumulation from water are dramatic; in soils, however, much of
the metal is bound to the soil material itself and must be mobilised into the soil
solution before it can be adsorbed. To achieve this the plant can secrete metalchelating molecules known as phytosiderophores which once in the rhizosphere
chelate and solubilise soil-bound metals. Two compounds, mugenic acid and
avenic acid behave as phytosiderophores in graminaceous species (Kinnersely,
1993). They are released from the roots in response to Fe and Zn deficiency and
mobilise Cu, Zn and Mn (Romheld, 1991). Plants can also solubilise metals by
lowering the pH of the soil environment, via the release of protons from the roots.
Citrate, an organic acid is secreted from plants and is able to chelate numerous
metals. It also plays a role in resistance to aluminium toxicity through its ability to
complex with Al(III) and prevent uptake by the plant (De la Fuente et al. 1997). It
is also possible to increase plant metal uptake by the addition of synthetic chelators
to the soil. Ethylene diamine tetra-acetic acid (EDTA) is a popular choice for
addition to lead contaminated soils, which can facilitate a 100 fold increase in the
uptake and transport of lead-EDTA-chelate complexes into stem and leaves
(Huang et al., 1997; Vassil et al., 1998). By increasing the secretion of appropriate
organic acids, it may be possible to increase uptake of metals.
Metal ions generally enter plant cells via specific or generic transmembrane
carriers or channels. Non-essential heavy metals can compete for the same
transmembrane carriers used by essential metals. This has been demonstrated with
Cu and Zn, essential metals, and Ni and Cd, non-essential metals (Clarkson &
Luttge, 1989).
Specialised carriers may also transport chelated metal complexes across the plasma
membrane. An example of this is Fe-phytosiderophore transport in graminaceous
species (Crowley et al., 1991).
More recent research has identified four proteins from Arabidopsis thaliana (a
model plant) - ZIP1, ZIP2, ZIP3 (Eng et al., 1998) and ITR1 (Eide et al., 1996) –
that all appear to be involved in facilitating pathways for the active uptake of toxic
metal ions, as a result of nutrient deficiency and stress. ZIP1 and ZIP3 mRNA has
been shown to accumulate in Arabidopsis roots as a result of zinc deficiency (Grotz
et al., 1998) and all contain an extra-membranal metal-binding motif HXHXH.
The ITR1 protein has been shown to transport both Cd(II) and Zn(II) (Yi &
Guerinot, 1996; Cohen et al., 1998). Once genes like these have been identified,
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they can be introduced into plants, using constitutive promoters enabling high
levels of expression, to examine whether or not metal uptake can be increased.
4.2.2 Transport within plants
Once metal ions have entered the root, they can either be stored or exported to the
shoot. Transport to the shoot probably occurs via the xylem, but may also occur
via the phloem (Stephan & Scholz, 1993). Movement of ions into the xylem may
be a rate limiting step because the metal ions must move from the cell cytoplasm to
the xylem by transversing the cell wall, symplastically. Xylem cell walls have a
high cation exchange capacity, therefore retarding the movement of cations. Noncationic chelated metal complexes such as Cd-citrate therefore, probably facilitate
the movement of metals via this route (Senden et al., 1992).
4.2.3 Hyperaccumulation
Some plants are able to concentrate metals in their tissues to levels far exceeding
those present in the surrounding soil or by other non-accumulating plants. A
generally accepted definition of a hyperaccumulator is one that is able to
accumulate more than 0.1% of Ni, Co, Cu, Cr and Pb or 1% of Zn in its leaves (dry
weight) (Baker, 1999). At these levels it has been proposed that the recovery of
metals would be economical (Baker, 1999), however, for the purposes of land
remediation accumulation of metal pollutants does not need to be so high.
Hyperaccumulators can be divided into groups according to the metals they
accumulate:
i)
Cu/Co accumulators
ii)
Zn/Cd/Pb accumulators and
iii)
Ni accumulators (Raskin et al., 1994).
Mostly they are restricted to only a few geographical locations and tend to be small
plants, not suited to biomass production. The majority of hyperaccumulators are
found in the cabbage family, Brassicaceae (Baker et al., 1994). One of the best
examples of a hyperaccumulator is Alyssum lesbiacum. It is able to accumulate
Ni(II) in the shoots and leaves to >3% of the dry weight (Krämer et al., 1996).
When this plant is exposed to nickel, a large and proportional increase in the level
of free histidine occurs which can be shown to be coordinated with nickel in vivo.
The molecular mechanisms involved in hyperaccumulation are poorly understood,
however when the genes involved are identified and isolated, their over-expression
in biomass crops could lead to great enhancement in metal tolerance and
accumulation.
Plants are able to protect themselves from the damaging effects of metal ions by
compartmentalization, chelation and precipitation of metal complexes. The two
most extensively studied systems employed by plants to achieve this, involve
cysteine-rich peptides; metallothioneins and phytochelatins.
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4.2.4 Metallothioneins
Metallothioneins (MTs) are low molecular weight cysteine-rich metal-binding
peptides. They have been found in animals, fungi, cyanobacteria and higher plants.
They are thought to provide housekeeping functions for the regulation of
intracellular concentrations of metals, ensuring, for example, that enzymes are
supplied with appropriate metal ions. In animals they have been shown to bind
copper, cadmium, zinc and silver and to detoxify normally lethal concentrations of
cadmium and copper in yeast (Tohayama et al., 1995). Numerous MT
genes/proteins have been identified in plants; wheat (Kawashima et al., 1992),
maize (De Framond, 1991), barley (Okumura et al., 1991), pea (Evans et al., 1990)
and soybean (Kawashima et al., 1991). A whole family of MT genes has been
identified in Arabidopsis thaliana (Zhou & Goldsbrough, 1995). The plant MTs
are usually 60-80 amino acids long with around 9-16 cysteine residues. Several
experiments have been carried to over-express various MT genes in plants in the
hope of conferring increased heavy metal tolerance and accumulation. One of the
first experiments of this nature to be reported is that of Misra & Gedamu, (1989).
They demonstrated that transgenic Brassica napus L. and tobacco, containing a
human metallothionein gene, were able to grow unaffected on 100M CdCl2, in
contrast to control seedlings which displayed severe inhibition of shoot and root
growth and chlorosis of leaves. In 1993 Pan et al. reported the expression of a
mouse metallothionein gene in transgenic tobacco plants. The MT gene was fused
to the cauliflower mosaic 35S (CaMV 35S) promoter which facilitates constitutive
expression of the gene in all cell types. Of the plants recovered from the
transformation, 20% displayed high levels of expression from the inserted gene and
were able to grow unaffected on media containing 200 M cadmium, whereas the
health of control plants was severely affected on just 10 M cadmium. Much work
has been done to characterise MT genes. They appear varied in the metals they
bind and in the manner with which their expression is regulated (Robinson et al.
1993). Further research to examine their potential for phytoremediation is
required.
4.2.5 Phytochelatins
Phytochelatins (PCs) are cysteine-rich proteins that have the ability to complex
metals (Ag(I), Hg(II), Pb(II), Zn(II), Cd(II) and Cu(II)) and sequester them in the
vacuole. They have been found in most higher plants, algae and in some yeasts.
PCs are synthesised via an enzymatic pathway that converts cysteine to glutamylcysteine, -glutamylcysteine to glutathione and glutathione to
phytochelatin. The three enzymes catalysing these steps are -glutamylcysteine
synthetase (-ECS), glutathione synthetase (GSH synthetase) and phytochelatin
synthase respectively. Although involved in PC synthesis glutathione (GSH) also
plays an important role in countering oxidative stress and may be conjugated to
many different xenobiotic compounds to facilitate their sequestration in the
vacuole. PCs are synthesised in response to heavy metals and can cause both a
reduction in the amount of GSH in the cell (Grill et al., 1987; Meuwley & Rauser,
1992) and an increase in GSH synthetase activity (Schneider & Bergmann, 1995).
Arabidopsis plants mutant in the synthesis of PCs or glutathione have been shown
to be hypersensitive to cadmium (Howden et al., 1995a, 1995b) demonstrating the
role of PCs in protecting plants from cadmium and possibly other toxic metals.
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By over-expressing the enzymes involved in the pathway, cellular levels of
phytochelatin might be increased with a corresponding increase in metal
sequestration. Two of the genes, -glutamylcysteine synthetase and glutathione
synthetase, in the pathway have been identified and cloned from several organisms
including Arabidopsis thaliana (May & Leaver, 1994; Rawlins et al., 1995) and E.
coli (Gushima et al., 1984; Watanabe et al., 1986). A further gene, HMT1,
isolated from the fission yeast Schizosaccharomyces pombe, is thought to be
involved in the compartmentalisation of heavy metals in the vacuole (Ortiz et al.,
1992). The deduced protein sequence of the gene displays homology with ABC
(ATP-binding cassette) type transport proteins suggesting that it may be an integral
membrane protein. It has also been found associated with vacuolar membranes
(Ortiz et al., 1995). Over-expression of the HMT1 gene in yeast confers enhanced
cadmium tolerance and increased levels of intracellular cadmium. However,
attempts to over-express the yeast gene in plants have been unsuccessful so far
(Ow, 1993). Whether homologues from plants can be identified remains to be
seen. Most recently, plant, animal and fungal genes encoding phytochelatin
synthase have been identified (Clemens et al., 1999; Vatamaniuk et al., 1999).
Exposure to cadmium has been shown to increase the synthesis of phytochelatin
synthase several fold and over-expression of plant phytochelatin synthase in yeast
increases tolerance and accumulation of cadmium (Vatamaniuk et al., 1999).
Two rate-limiting steps have been identified in PC synthesis; the supply of cysteine
and the step catalysed by -ECS. This was deduced from experimental evidence
resulting from the over-expression of -ECS in transgenic poplar (Noctor et al.,
1996). Transgenic plants over-expressing -ECS were found to contain 10 and 3
times more -glutamylcysteine and glutathione respectively and the application of
exogenous cysteine increased the glutathione content further in both transformed
and untransformed control plants. Based on the evidence from Noctor et al., 1996
it would be useful to examine whether or not increased levels of PCs are obtained
as a result of over-expressing -ECS and whether the poplar plants are able to
tolerate and accumulate higher levels of metals than non-transformed control
plants. In addition to the above, glutathione synthetase is also suggested as being
rate limiting. When the E. coli glutathione synthetase gene was over-expressed in
Brassica juncia, transformed plants were found to have higher concentrations of
GSH and PC and increased tolerance to cadmium than the control plants (Zhu et
al., 1999). It appears therefore, that the manipulation of glutathione and
phytochelatin concentrations offers good potential for increasing metal tolerance
and accumulation in plants.
4.2.6 Transformation of toxic elements
The transformation of toxic elements into relatively non-toxic forms using plants
has been demonstrated using transgenic Arabidopsis thaliana to convert organic
mercury to much less toxic elemental mercury (Bizily et al., 2000). The
Arabidopsis thaliana plants were transformed with two bacterial genes:
(i)
merA which encodes mercuric reductase and
(ii)
merB encoding organomercurial lyase.
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BIOENERGY CROPS AND BIOREMEDIATION
These genes have been cloned from bacteria isolated from organic-mercury
contaminated soils. The enzymes they encode are capable of converting
methylmercury and other organomercurials to elemental mercury, which is volatile
and diffuses out of the bacterial cells. Arabidopsis plants expressing both of the
genes were found to grow on 50 times higher concentrations of methylmercury
than control plants and were shown to transpire elemental mercury. This work
demonstrates the potential for phytoremediation using genes from other organisms.
It now requires scaling up from the model plant, Arabidopsis thaliana, to a
biomass crop followed by field trials.
Selenium is a particular problem in the wetlands of western USA. Transgenic
expression of the Arabidopsis plastid ATP sulfurylase (APS1) in Brassica juncia
has been shown to increase the uptake of selenate and its reduction to selenite
(Pilon-Smits et al., 1999). Plants were also more able to tolerate selenate. It has
also been shown that the volatilization of selenium by plants is enhanced by
bacterial activity in the rhizosphere (De Souza et al., 1999). It is thought that many
phytoremediation processes are enhanced in this way, which suggests that further
research examining plant/bacterial co-existence would be of value.
The major form of iron found in soils is Fe(III), which is relatively unavailable to
plants and fairly toxic. Arabidopsis thaliana is an example of a plant that is able to
reduce Fe(III) to Fe(II) fairly well. It is able to do this via a ferric chelate reductase
(FRO2) at the root surface (Yi & Guerinot, 1996; Robinson et al., 1999). Fe(II) is
then transported into the root cells by a ZIP transporter (discussed earlier). Plants
that are mutant for FRO2 are iron deficient, but over-accumulate Cu(II), Zn(II) and
Mn(II) as they try to extract more iron from the soil. However, Arabidopsis
mutants for a manganese accumulator, show co-ordinate deregulation for Fe(II),
Cu(II), Zn(II), Mn(II) and Mg(II) and also for FRO2 activity (Delhaize, 1996).
4.3
PHYTOREMEDIATION OF ORGANIC POLLUTANTS
Strategies employing the use of microorganisms for the detoxification of organic
waste products have been used for many years. More recently however,
microorganisms, that have evolved catabolic genes enabling them to detoxify toxic
chemicals, have been utilised for the bioremediation of polluted land. Using
natural bacterial gene transfer systems it has been possible to transfer certain
degradative processes to certain bacteria, increasing the potential for remediation
under differing circumstances. Recent advances in molecular biology have taken
this concept further by making it possible, in theory, to introduce virtually any
gene into any organism. However, experiments to evaluate the effectiveness of
bio-augmentation, the addition of microorganisms for the purposes of remediation,
have been disappointing. The reasons for this include:
(i)
competition from endogenous microorganisms causing rapid decreases in
populations of introduced microorganisms and
(ii)
there is often a need for induction of the degradative genes by the
application of undesirable inducers to the land.
The biodegradative capability of plants is much less impressive than those of
adapted bacteria and fungi. The alternative, therefore, is to use the array of
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catabolic bacterial genes for expression in biomass plants, combining the best
elements from two systems for effective remediation. In this way genes can be
designed without the need for induction using undesirable compounds.
The mechanisms controlling the uptake and sequestration of toxic organic
compounds in plants are relatively unknown. One system we do know something
about is the pumping out of the cell or sequestration into the vacuole of glutathione
conjugates of organics and other conjugated-toxic organic complexes. This is
facilitated by a glutathione-S-conjugate pump, the gene for which has been isolated
from Arabidopsis thaliana (Tommasini et al., 1998).
Halogenated compounds like TCE are difficult to metabolize and are generally
toxic and carcinogenic. They are widespread contaminants and pose significant
risks to animals and humans. Plants are known to produce aliphatic dehalogenases
that degrade TCE (Anderson et al., 1993; Anderson & Walton, 1996). They have
been shown to extract TCE from polluted sites and transpire it. Root exudate has
been shown to enhance degradation of TCE in the rhizosphere. Axenically grown
hybrid poplar has been demonstrated to remove TCE from the surrounding
medium and convert it to trichloroethanol, chlorinated acetates and finally CO2
(Gordon et al., 1998). In one experiment, using poplar cells in tissue culture, more
than 10% of the TCE present was converted to CO2 within 10 days. More recently,
TCE remediation has been investigated using transgenic plants expressing a
mammalian cytochrome P450 2E1 (Doty et al., 2000). This enzyme oxidises many
compounds including; TCE, dibromoethane (EDB), carbon tetrachloride, benzene,
styrene, chloroform, 1,2-dichloropropane and vinyl chloride. Transgenic tobacco
plants expressing the introduced gene displayed a 640-fold increase in the
metabolism of TCE as compared to control plants. Transgenic plants also
displayed increased uptake and debromination of ethylene dibromide.
Cytochrome P450 monooxygenases have been found in plants, bacteria and
animals and catalyse many types of chemical transformations, including aliphatic
hydroxylations, expoxidations, dealkylations, dehalogenations and many
deactivations (Guengerich & MacDonald, 1990). Several hundred P450 enzymes
have been characterised at the primary sequence level and other natural P450s have
yet to be isolated and characterised. In addition to the natural P450s available,
much work is being carried to specifically alter substrate specificity and catalytic
efficiency of P450s using protein engineering (Kellner et al., 1997). The ability to
degrade xenobiotics and their wide range of substrate specificities makes
cytochrome P450s very good candidates for exploitation in phytoremediation.
A combination of bacterial gene in a plant system has been tested to evaluate
whether or not explosive residues could be remediated. A gene encoding
pentaerythritol tetranitrate reductase, isolated from an explosive-degrading
bacterium, was introduced into tobacco. Seeds from the transgenic plants
expressing pentaerythritol tetranitrate reductase were able to germinate and grow
on 1 mM glycerol trinitrate (GTN) or 0.05 mM trinitrotoluene; concentrations that
inhibited the growth of control seeds (French et al., 1999). Transgenic seedlings
grown in liquid medium containing 1 mM GTN displayed more rapid and complete
denitration of GTN than control seedlings. Again, these results offer promise for
the enhancement of remediation by plants.
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4.4
PLANT TRANSFORMATION
The availability of appropriate genes and strategies for the over-expression of these
in biomass crops, relies on the ability to transfer such genes to the target plant in
such a way that a copy of the foreign DNA is stably integrated into every cell of
the plant. To achieve this, it is necessary to have a tissue culture system, specific
to the plant, that enables the regeneration of whole plants from single totipotent
cells. Once a successful tissue culture system has been developed, a plant
transformation system must be established to enable the transfer of foreign DNA to
those single cells capable of regeneration into whole plants. It is widely recognised
that monocotyledonous plants are generally more difficult to transform than
dicotolyedonous plants - and grasses (Poaceae) in particular have proven
problematic.
The totipotent cells can be generated either as embryogenic suspension cultures or
as embryogenic callus tissue on solid media. Plant cell suspension cultures able to
regenerate plants through somatic embryogenesis were first reported in carrot in
1958 (Steward et al., 1958). It was not until the early 1980s that embryogenic
suspension cultures were achieved for species from the Poaceae family, of which
Miscanthus is a member (Lu & Vasil, 1981; Vasil & Vasil, 1981; Ho & Vasil,
1983). Efficient embryogenic suspension culture systems have now been
established for the most commercially important members of the Poaceae family,
including Miscanthus (Holme, 1996). Callus induction and plant regeneration on
solid media has also been achieved using various explant types from Miscanthus x
ogiformis Honda ‘Giganteus’ (Holme & Petersen, 1996; Petersen et al., 1999). We
are not aware of any published literature describing the stable transformation of
Miscanthus tissues, however, our own unpublished work has shown that
Miscanthus callus tissue can be transformed via microprojectile bombardment
resulting in transient expression of a foreign reporter gene, GUS, driven by a
ubiquitin promoter. It is not unreasonable to expect that the development of a
stable transformation system for Miscanthus is underway. Combined with the
extensive work, carried out at the Danish Institute of Plant and Soil Science, to
establish a successful tissue culture system, it is likely that the generation of
transgenic Miscanthus will not be long in the making.
In contrast to Miscanthus, poplar species have received much greater attention with
respect to the introduction and over-expression of foreign genes. The first reports
of poplar transformation were published in the late 1980s (Parsons et al., 1986;
Fillatti et al., 1987). Since then, a number of different poplars have been
successfully transformed using the two most popular transformation methods
available; microprojectile bombardment and Agrobacterium tumefaciens-mediated
transformation. These include, amongst others, white poplar (Populus alba)
(Confalonieri et al., 2000); the hybrid poplars, Populus tremula x alba (Gallardo et
al., 1999; Franke et al., 2000) and Populus x canadensis (Confalonieri et al., 1997;
Liang et al., 1999); and eleven different hybrid cottonwood genotypes, several of
which were previously difficult to transform and are economically important (Han
et al., 2000).
Surprisingly, willow appears to have received little attention with respect to the
development of tissue culture or plant transformation systems. Amo-Marco &
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Lledo (1996) report the development of in vitro culture techniques for the
propagation of an endangered willow, Salix tarraconensis. In addition, Xing &
Maynard (1995) report the production of transgenic shining willow (Salix lucida).
In this experiment, shoot explants were transformed using Agrobacteriummediated transformation, the introduced gene encoding -glucuronidase, a protein
that can be used as a marker for gene expression. Willow is being examined for its
potential to phytoremediate land and water. The EC-funded BIORENEW project
is developing a rapid screening system to facilitate the assessment of metal
tolerance from large numbers of willow varieties (Watson et al., 1999).
Preliminary results suggest that plant responses tested in a hydroponic system
mimic responses seen under field conditions. Willow has also been shown to
facilitate the transformation of perchlorate into chloride (Nzengung et al., 1999).
Two phytoprocesses were identified as being responsible for decontamination:
(i)
uptake and phytodegradation in tree branches and leaves and
(ii)
rhizodegradation.
Although willow is being examined for phytoremediation purposes, it appears that
the step to enhancing phytoremediation processes using genetically modified
willow has not been taken. Progress in the development of transformation systems
for biomass willow is necessary to ensure that transgenic varieties can be
developed in the near future.
Transgene integrations are essentially random, and are therefore, subject to
position effect variation which affects the level of transgene expression. A general
approach to overcoming these limitations usually involves generating a large
number of transgenic plants, from which the ideal transformant can be selected.
This type of approach can be time consuming and costly and particularly difficult
when dealing with plants whose transformation efficiency is low. One method that
can help to alleviate positional effects is the use of matrix attachment regions
(MARs), which are lengths of DNA added to either side of the transgene and shield
the transgene from surrounding influence. This method has been used successfully
in poplar transformation, enhancing transformation frequency and transgene
expression (Han et al., 1997).
4.5
RISKS
In addition to the many potential benefits of genetic modification, it is recognised
that the introduction of foreign genes into plant genomes and the expression of
foreign proteins in plants have the potential to result in undesirable effects on the
environment and human health. The release of genetically modified organisms
(GMOs) into the environment, is, therefore, controlled in the UK by domestic and
European legislation as defined by Directive 90/220/EEC and implemented by the
Genetically Modified Organisms (Deliberate Release) Regulations 1992 (amended
1995 and 1997). This serves to prevent or minimise damage to the environment
via a statutory system of risk assessment and prior consent before any GMO may
be released or marketed. All applications are received by the Department for
Environment, Food and Rural Affairs (DEFRA), and the Advisory Committee on
Releases to the Environment (ACRE) advise on whether or not release should be
granted. This decision is arrived at, after careful consideration of the required
89
BIOENERGY CROPS AND BIOREMEDIATION
information supplied with each application. Central to this is an environmental
risk assessment that considers the potential risk to the environment and human
health posed by the release of the organism in question.
The accumulation of high levels of metals in plant tissues raises issues over the
toxicity of these plants to other organisms. It may be useful in eliminating pest
damage, for example, but could also adversely affect beneficial insects and other
organisms. The potential for the generation of unexpected toxic compounds during
the conversion of organic or other pollutants to less toxic forms is also an
important consideration that would need to be monitored closely. Changing just
one step in a metabolic process can have a dramatic effect on the production of
many different compounds. For these reasons it is important that the dispersal of
transgenes conferring pollutant remediation abilities should be minimised or
prevented. This can be achieved in a number of ways and is the subject of much
investigation (Advisory Committee on Releases to the Environment, 2000). The
most obvious solution might be to use plants that are genetically sterile. Sterility
can of course occur naturally, but can also be conventionally bred into suitable
plant lines and the trait can even be genetically engineered into plants. Poplar has
been genetically engineered for both sterility and accelerated flowering (Meilan
and Strauss, 1997). Another possible solution might be to insert the required DNA
sequences into the chloroplast genome rather than the nuclear genome. Most
plants inherit the chloroplast DNA maternally, avoiding dispersal in pollen.
4.6
DISCUSSION
The development of genetically modified plants for phytoremediation is still in its
infancy, but investigations to date suggest great potential for increasing the ability
of plants to remove pollutants from contaminated land and water. Many genes
involved in facilitating tolerance, accumulation and detoxification of metal and
organic pollutants have been identified in plants and other organisms, but there are
still a great many to be discovered. For many of the pathways facilitating the
detoxification of pollutants in plants, it will be possible to enhance detoxification
by modifying rate limiting steps by the over-expression of the corresponding
genes. In addition to modifying plant pathways, it has also been shown to be
possible to express genes from other plant species and even genes from animals,
bacteria, yeast and fungi. Much of this work has been performed in model plant
species such as Arabidopsis thaliana and tobacco. Poplar transformation has been
the subject of much study.
There are many publications reporting the
transformation of poplar species, including examples specifically investigating
phytoremediation potential. Other biomass crops have not been extensively
studied and therefore require further development.
Genetic engineering facilitates a choice of promoter to drive the expression of the
required gene. By choosing promoters that are induced by metals, organics or
other compounds, expression of the gene may be controlled on a demand basis. In
addition, the use of tissue-specific promoters may enable the accumulation of
metals in specific parts of the plant; for example, in the harvestable tissues, such as
the shoots and leaves, and not in the below ground tissues. It is possible to
introduce numerous genes to the plant genome during plant transformation and
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BIOENERGY CROPS AND BIOREMEDIATION
also to re-transform an already transgenic plant. This means that in instances
where several genes are required to facilitate the remediation of several pollutants,
a suitable plant can be generated. The most successful remediation strategies
might be those using a combination of genetically engineered or conventional
plants and associated rhizosphere bacteria.
Careful consideration must be given to the development of plants for
phytoremediation, because of the potential for adverse environmental effects. The
prevention of transgene flow to conventional plants and possible effects on other
organisms are important issues.
Dramatic improvement in phytoremediation capability has yet to be realised, but it
will happen as soon as more of the fundamentals of pollutant uptake, transport and
accumulation or transformation are elucidated.
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CHAPTER 5
LEGISLATION AND CODES OF PRACTICE AFFECTING THE
RECYCLING AND LAND APPLICATION OF ORGANIC WASTES
GORDON HICKMAN
5.1
INTRODUCTION
This chapter provides an overview of the main legislation and codes of practice
affecting the disposal and recycling to land of farmyard manures, sewage sludge
and other organic wastes that might be applied to biomass crops. It is based on
desk-based research and expert knowledge, but does not purport to be a definitive
listing of legislation which would impact on all wastes that might be used for
bioremediation. The main aim is to highlight the key issues and potential barriers.
5.2
UK WASTE REGULATIONS AND EC DIRECTIVES
The possible adverse environmental effects of agricultural activities have been a
concern in many European countries since the mid-1970's. Concerns were initially
related to water pollution and odours, but more recently have included emissions of
ammonia and greenhouse gases and protection of the soil.
In 1999, agriculture caused 4,254 (14% of the total) substantiated water pollution
incidents in England and Wales (EA, 2001). Of these, some 238 (24% total) were
regarded as major or significant incidents (category 1 & 2). Although this figure
includes incidents due to pesticides, fuel oil and other contaminants, 2,161 (7%
total) substantiated incidents were caused by livestock excreta, silage effluent or
other organic wastes such as sewage sludge or raw sewage. Legislation has often
been aimed at reducing such point source water pollution incidents.
Table 5.1 summarises some of the key legislation relating to the actual disposal or
recycling to land of organic wastes. Further commentary is provided in the
following sections.
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BIOENERGY CROPS AND BIOREMEDIATION
Table 5.1. Summary of legislation and voluntary codes of practice affecting
disposal of farm manures and organic wastes.
Title
Type of waste affected
What does it regulate?
Control of Pollution
Regulations 1991/1997
Manures, slurry, silage
effluent and fuel
New storage structures.
Requirements for
impermeability, structural
stability and siting in relation to
watercourses.
Nitrate Vulnerable Zone
Regs. (EC Nitrate
Directive)
Animal manures,
sludges, wastes high in
available N
Manure N application rate and
timing.
MAFF/WOAD Water
Code 1998
All types
All aspects of avoiding pollution
from manures, including odours,
ammonia and gaseous
emissions.
Waste Management
Licensing Regs, 1994 (EC
Framework Directive on
Waste), HMSO, 1994.
Industrial, household
and commercial wastes “controlled wastes”
Licensing requirements and
Duty of Care. (Beneficial
wastes exempted subject to
limits on application rate).
Integrated Pollution
Prevention and Control
(IPPC) Regs. (2000).
(EC Directive 96/1)
All types
All types of emissions. To
apply from 2004 to disposal
sites for non-hazardous wastes,
food and drink installations from
2004 (typically) and from 2007
to units for the intensive rearing
of poultry or pigs (with >40,000
poultry, 2000 pig places, or 750
sow places).
Sludge (Use in
Agriculture) Regs. (1989)
and Code of Practice for
the Agricultural Use of
Sewage Sludge (1996)
(EC Sewage Sludge
Directive 86/278/EEC)
(Currently under revision)
Sewage sludge and
sludge products
Application rates and timing,
management of sludge; limits on
soil metal levels and metal
additions.
“Safe Sludge Matrix”
Voluntary agreement
currently being
incorporated into Sludge
(Use in Agriculture) Regs.
Sewage sludge of all
types
Guidance on minimum
standards for application
practice and minimum
acceptable level of treatment for
any sludge product (biosolids) to
be applied to any crop or
rotation.
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BIOENERGY CROPS AND BIOREMEDIATION
5.2.1 Codes of practice
The Department for Environment, Food and Rural Affairs (DEFRA), formerly
MAFF, and National Assembly for Wales Agriculture Department (NAWAD),
formerly WOAD, publish three separate Codes of Good Agricultural Practice for
the Protection of Water, Air and Soil - which give guidance on ‘best practice’ for
avoiding pollution. These Codes were first published in 1991-3 and were revised
and re-issued in 1998 (MAFF, 1998a, 1998b & 1998c).
These Codes are statutory codes under Section 97 of the Water Resources Act 1991
(HMSO, 1991). This means that although failure to comply with the Codes is not
an offence, it would be taken into account in any legal action taken as a result of a
pollution incident. Furthermore, many Farm Assurance Schemes for both crops
and livestock include compliance with the codes as part of their protocols. All the
Water Operators have also signed up to complying with these codes.
The Codes also detail the limits to application rates for organic materials, namely
250 kg ha-1 total N per year, or a single application of 500 kg ha-1 total N of solid
materials with limited N availability in non sensitive catchments every two years.
The Codes also give guidance on the identification of areas where manures should
not be spread, or where rates or timing should be restricted to minimise the risk of
surface runoff. A 10 m wide no-spreading zone is recommended adjacent to
watercourses and a minimum of 50 m radius adjacent to springs, wells or boreholes
supplying water. Other areas can be classified for water pollution risk according to
slope, soil characteristics and proximity to watercourses. Farmers are encouraged
to produce a farm waste management plan, which takes account of these factors.
Such plans normally incorporate a map of the farm with field areas annotated in
different colours, to indicate restrictions on timing of spreading or amounts to be
spread.
The Water Code also differentiates between low and high available N organic
manures, viz.:
Low available N High available N -
sludge cake, thermally dried and lime-treated sludge
(straw-based farmyard manures)
liquid digested sludge (cattle/pig slurry and poultry
manures)
The Code also recommends that where practically possible, manures containing a
high proportion of available N should not be applied to arable land in the autumnearly winter period, to minimise the risk of nitrate leaching.
Impact and implications for bioremediation
The Codes of Practice provide a framework of guidance within which most
operators work when applying organic wastes to agricultural land. They are not
likely to limit or hinder the use of wastes on bioenergy crops for bioremediation.
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BIOENERGY CROPS AND BIOREMEDIATION
5.2.2 Odours and gaseous emissions
General guidance and legislation
Justifiable complaints from the public about odours from agriculture totaled 66,658
in 1994. Half of these were due to land spreading of manures and the majority of
complaints concerned odours arising from pig farms. Guidance on minimising
odours from buildings, manure stores and land spreading is given in the Code of
Good Agricultural Practice (COGAP) for the Protection of Air. Under the
Environmental Protection Act 1990, an abatement notice can be served on the
owner of a unit, which causes a statutory nuisance, which includes odours (HMSO,
1990a).
Integrated Pollution Prevention and Control (IPPC) legislation.
The system of IPPC comes under the Pollution Prevention and Control (England
& Wales) Regulations 2000, as amended (SI, 2002). It is a means of applying an
integrated environmental approach to the regulation of certain industrial activities.
This means that emissions to air, water and land, plus a range of other
environmental effects, must be considered together. IPPC aims to prevent
emissions and waste production and, where that is not practicable, to reduce them
to acceptable levels. Waste reduction should be treated as a priority and energy
must be used efficiently.
Installations falling under the requirements of IPPC will be given permits to
operate, providing certain conditions are met. These permits will require that best
available techniques (BAT) are used in meeting the requirements of the conditions.
In order to apply for a permit under IPPC, the applicant must provide detailed
information regarding the operation of the installation in question. For example, an
intensive livestock farmer will be expected to supply, amongst other information,
the following:





Details of the management structure and staffing arrangements, including
details of training and emergency planning.
A raw materials inventory.
Records of current water and energy use.
Confirmation that a water and energy audit will be undertaken within 18
months of permit issue date.
Manure management plans.
Under the PPC Regulations, installations fall under one of three classifications
dependant upon the activities that take place within them:
 The Environment Agency regulates Part A(1) installations.
 Part A(2) installations are regulated by the relevant local authority.
However, the local authority will always be the statutory consultee where
the Environment Agency is the regulator and vice versa. The two will
work together in the permitting process.
 Part B installations come under the regime of Local Air Pollution
Prevention and Control (LAPPC). LAPPC is similar to IPPC from a
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BIOENERGY CROPS AND BIOREMEDIATION
procedural point of view but focuses on controlling air emissions only.
Local authorities are the regulators.
Impact and implications for bioremediation
The impacts are difficult to assess, since the implementation timetable is still some
way off. The regulations should result in less waste being produced, as well as
more responsible and informed disposal/recycling of any waste that is produced.
The inclusion of disposal sites treating more than 50 tonnes/day of non-hazardous
wastes, such as food and livestock waste, by biological or physio-chemical means
may provide some barriers to development of composting and anaerobic digestion
sites. However, they are not likely to limit or hinder the use of wastes on
bioenergy crops for bioremediation, unless it was decided that the BAT for
application of liquids was injection. If this were the case, it would limit
applications to pre-establishment in the case of SRC and possibly Miscanthus.
5.2.3 Water pollution
General
Under the Water Resources Act 1991 (HMSO, 1991) it is an offence to cause
pollution of groundwater, lakes, ponds, rivers, streams and canals. Farmers and
others may be prosecuted for causing such pollution and fined up to £20,000 in a
local court. In addition the Act contains provisions for secondary legislation which
aims to prevent pollution occurring in the first instance.
Storage of Slurry and Manure
The Control of Pollution (Silage, Slurry and Agricultural Fuel Oil) Regulations
1991 (SI, 1991), introduced minimum requirements for the size and siting of new
or substantially modified slurry storage structures and laid down minimum
constructional standards based on BS 5502 (Buildings and Structures for
Agriculture).
No specific controls exist for solid manures temporarily stored in field areas
without a structured base, although there are recommendations on siting. The EA
is able to regulate and control the use of long-term field stockpiles.
Groundwater Protection Zones
Up to 35% of public drinking water is derived from groundwater sources such as
aquifers, and the protection of such sources is an important function of the EA.
Source protection zones have been established throughout England and Wales and
in most cases waste materials (does not currently apply to farm wastes) are not
permitted in Zone 1 (most sensitive) catchment areas or around private boreholes.
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BIOENERGY CROPS AND BIOREMEDIATION
Impact and Implications for Bioremediation
The regulations provide a framework of guidance, within which most operators
work when applying organic wastes to agricultural land. They are not likely to
limit or hinder the use of wastes or bioenergy crops for bioremediation.
5.2.4 Land application,
materials
including
bioremediation,
The recycling of nutrients in livestock manures is excluded
regulations on the use of ‘waste materials’ spread to land.
specified geographical areas, such as in Nitrate Vulnerable
Special Scientific Interest (SSSIs), there are no regulations
timing or rate of application of manures.
of
organic
from EC and UK
Except in certain
Zones or Sites of
which govern the
DEFRA have recently updated and revised its guidance on fertiliser use in
Fertiliser Recommendations for Agricultural and Horticultural Crops (RB209) 7th
Edition (2000). Although primarily aimed at providing guidance on giving the best
financial return to farmers, the guidelines should also help to minimise nutrient
losses to the water and air.
Recommendations, which encourage the optimum use of nutrients in manures, also
help to minimise diffuse water pollution risks from land spreading. Literature, like
booklet RB209, is available to give typical analyses of manures and slurries and
details of how to calculate optimum application rates for particular crops, soils and
application methods. Computer programs to aid decision-making, such as
MANNER (ADAS, 2000), are also available. Whilst farmers are encouraged to
take full account of the nutrients available in manures when planning their fertiliser
applications, many do not.
Impact and implications for bioremediation
The codes of practice and guidance provide a framework within which most
operators work when applying organic wastes to agricultural land. They are not
likely to limit or hinder the use of wastes on bioenergy crops for bioremediation.
There is, however, relatively little available information on nutrient requirements
of bioenergy crops under a range of soil types and climatic conditions; although the
apparently low requirement of some may limit the use of ‘nutrient rich’ organic
wastes, if application rates are limited to probable crop uptake.
5.2.5 Draft Soil Strategy – England
DEFRA has published a consultation draft of a Soil Strategy for England (DETR,
2001). The document was produced as a direct response to recommendations
made by the Royal Commission on Environmental Pollution. The overall strategy
is to ensure that soil is used and protected in such a way that it is sustainable in its
own right and that it contributes to sustainable development generally.
One of the key objectives is to address the issue of soil erosion, through R&D on
the extent and causes of soil erosion on agricultural land. It also refers to the
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BIOENERGY CROPS AND BIOREMEDIATION
promotion of good practice for the land treatment of industrial wastes and the
extension of waste management controls to agricultural wastes.
Impact and implications for bioremediation
The strategy is still in a draft form but gives an indication as to the likely direction
of future policy. The proposals will be incorporated into the relevant legislation
and codes of practice discussed elsewhere in this section and are not likely to limit
or hinder the use of wastes on bioenergy crops for bioremediation. The possible
inclusion of farm wastes under waste management controls may have an impact,
depending on how the regulations are introduced. The focus on erosion could
present some limitations to bioenergy crops, since there is some limited evidence
to suggest that the reduced ground cover and relatively slow establishment of short
rotation coppice and other bioenergy crops can lead to increased risk of erosion on
some sites.
5.2.6 Nitrate
The UK government has established 32 Nitrate Sensitive Areas (NSAs), where
farmers are paid compensation to voluntarily join a scheme which aims to reduce
nitrate loss from the soil to groundwater. In these areas there is a maximum limit
of 250 kg total N ha-1 in any 12 month period from applied organic manures,
limitations on timing of slurry and poultry manure applications and various other
control options affecting the way in which farmers may use fertilisers and their
land. The rules have been successful in reducing nitrate loss from the soil zone.
Under the EC Nitrate Directive (EC, 1991b) 68 Nitrate Vulnerable Zones (NVZs)
in England and Wales, totalling 600,000 ha, were designated in April 1996; where
compulsory limits on the application, and deposition by grazing animals, of total
nitrogen are enforced. An upper limit of 250 kg ha-1 yr-1 of organic N is allowed
on grassland and 210 kg ha-1 on arable land [reducing to 170 kg ha-1 yr-1 after four
years (from 19 December 2002)]. In addition, the application of slurries and
poultry manures are banned on sandy and shallow soils during the late summer and
early autumn, when there is little crop growth removing available nitrogen from
the soil. Within Nitrate Vulnerable Zones, the rules are compulsory and no
compensation is paid to farmers, as they are regarded as ‘good agricultural
practice’.
Impact and implications for bioremediation
The regulations provide a framework of guidance within which most operators
work when applying organic wastes to agricultural land. They are not likely to
limit or hinder the use of bulky solid wastes on bioenergy crops for bioremediation,
but the ‘closed periods’ limit the use of slurries and liquid wastes in the autumn.
Biennial cropping of short rotation coppice (instead of the usual three-yearly
harvesting) may provide an opportunity to take advantage of the option for
‘double’ applications, of 500 kg ha-1 total N, every two years. Applications could
be made in late winter or spring, after harvest. Such applications would not,
however, be permitted in NVZs, where there will be an annual limit of 250 kg ha-1
total N from organic sources.
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The number of NVZs is expected to significantly increase in the near future and
depending on the extent of the new areas covered, may impose some additional
constraints on the use of organic wastes in these areas.
5.2.7 Phosphate
The Environment Agency has identified nutrient enrichment as one of the ten
priority issues requiring attention, in order to achieve a more sustainable balance
between the needs of society and the needs of freshwater ecosystems. The issue of
phosphate enrichment of water is becoming increasingly important, given the link
between phosphate and eutrophication of lakes, reservoirs and slow moving rivers.
The Environment Agency has identified agriculture as being a particular problem,
since most of the pollution occurs via diffuse pollution. The Environment Agency
estimates that livestock (34% of phosphates), human and household waste (24%),
fertiliser (16%) and detergents (10%) represent the major sources of phosphate
inputs to surface waters.
The Soil and Water Codes (MAFF, 1998a & b) provide guidance on limiting
phosphate losses and recommend that, at soil index 3 and above, care should be
taken to avoid total phosphorous inputs exceeding the amount removed by crops in
the rotation. The guidance is however not statutory and the Environment Agency
have stated that unless inputs can be controlled by voluntary means, regulatory
restrictions may be required.
Some restrictions are possible under the provisions of the Urban Wastewater
Directive (EC, 1990). The UK has designated 80 Sensitive Areas (Eutrophic)
covering both running and standing water. The number and size of such areas may
be increased in the future. There are some 140 sites currently being monitored and
a review of areas is to be undertaken during 2001.
The Environment Agency has stated that it does not intend to apply for powers to
establish water protection zones under Section 93 of the Water Resources Act
(HMSO, 1991), although it will do so in catchments where other control initiatives
have failed.
Impact and implications for bioremediation
The issue of phosphate will undoubtedly become even more important and may
eventually give rise to the creation of ‘phosphate sensitive zones’. These could
have an effect on high phosphate materials such as sewage sludge. However, the
current regulations and codes are not likely to hinder the use of wastes on
bioenergy crops for bioremediation - although the apparently low phosphate
requirements of bioenergy crops could limit the total amount of organic matter that
can be applied to soils, since the limiting constituent may, in some cases, be the
total phosphate content.
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5.2.8 Sewage sludge
Sewage sludge applications to land are regulated by European Community
Directive 86/278/EEC (EC, 1986) which has been implemented in England,
Scotland and Wales by The Sludge (Use in Agriculture) Regulations 1989 (SI,
1989) and as amended by The Sludge (Use in Agriculture) Amendment Regulations
1990 (SI, 1990). These are complemented by the DEFRA Code of Practice for
Agricultural Use of Sewage Sludge (DoE, 1996) and Codes of Good Agricultural
Practice for the Protection of Water (MAFF, 1998a), Air (1998b) and Soil (1998c).
In operational practice, sludge use on agricultural land needs to comply with the
DEFRA Water Code recommendation that applications should not supply more
than 250 kg total N ha-1 per annum, which in most situations will provide the
working limit on major nutrient and heavy metal loading rates. The revised
DEFRA Water Code (MAFF, 1998a) does, however, allow low available N
manures, such as composts and sludge cakes, to be applied at 500 kg total N ha-1
every other year.
The above regulations are currently being revised and a consultation draft is
expected later in 2001. The regulations will adopt the provisions of the Safe
Sludge Matrix and for the first time introduce standards for pathogen content in
sludges. The regulations will also enforce a ban on the use of untreated sludges on
food crops, and will include septic tank liquor within the regulations as untreated
sludge – thereby preventing its application to grassland and other food crops. (The
Waste Management Licensing Regulations will also be amended accordingly).
Impact and implications for bioremediation
The regulations provide a framework of guidance within which all operators work
when applying sewage to agricultural land. They are not likely to limit or hinder
the use of wastes on bioenergy crops for bioremediation. The revised regulations
will however ban the use of untreated sewage sludge to all crops from 31
December 2005.
5.2.9 Safe Sludge Matrix
Negotiations involving the UK Water Industry, the British Retail Consortium
(BRC) representing the major retailers and ADAS, aimed to secure a sustainable
route for recycling sludge to agricultural land that was acceptable to the food
industry, water industry, regulators and farmers and growers were concluded in
July 1998.
The ‘Safe Sludge Matrix’ agreement (ADAS, 2001a), commonly known as the
ADAS Matrix, came into being on 31 December 1998, and has been accepted as the
minimum standard for sustainable sludge recycling to agricultural land. The
Matrix consists of a table of crop types, together with clear guidance on the
minimum acceptable level of treatment for any sludge based product (commonly
referred to as biosolids) which may be applied to that crop or rotation (Table 5.2).
All UK outdoor crops are covered from grass for grazing and silage making, maize
for silage, combinable crops and animal feed crops, through to horticultural crops,
vegetables, salads and fruit.
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BIOENERGY CROPS AND BIOREMEDIATION
Since December 1999, untreated sludge has no longer been allowed on land used to
grow food crops although, under the agreement, industrial crops grown under
contract for non-food use (e.g. industrial crops grown on set-aside) continued to
receive untreated sludges until December 2001. The surface spreading of treated
sludge on grazed grassland was banned from 31 December 1998. Treated sludge
can only be applied to grazed grassland where it is deep injected. More stringent
treatment processes are required where sludge is applied to land growing vegetable
crops and in particular those crops that may be eaten raw (e.g. salad crops).
Treated sludge can be applied to agricultural land, which is used to grow
vegetables - provided that at least 12 months have elapsed between application and
harvest of the following field vegetable crop. Where the crop is a salad, which
might be eaten raw, the harvest interval must be at least 30 months.
Table 5.2. The ‘Safe Sludge Matrix’
“Safe Sludge Matrix”
Crop Group
Fruit
Salads
Untreated
sludges
x
x
Treated
sludges
Enhanced treated
sludges
x
x


(30 month harvest
interval applies)
Vegetables
x
x

(12 month harvest
interval applies)
Horticulture
Combinable
& Animal
Feed Crops
Grass - GRAZED
&
Forage- HARVESTED

x
x

x



x
x
(Deep injected or
ploughed down only)
x

10
month
harvest
interval
applies)
3 week
no grazing
and
harvest
interval
applies



3 week
no grazing
and
harvest
interval
applies
All applications must comply with the Sludge (Use in Agriculture) Regulations 1989
 and
DoE Code of Practice 1996
x Applications not allowed (except where stated conditions apply)
An ‘Enhanced Treated Sludge’ (previously known as ‘Advanced’) category has
been included in the Matrix, to describe treatment processes that are capable of
virtually eliminating any pathogens that may be present in the original sludge (e.g.
thermal drying). This is the first time in the UK that there has been recognition of
differences in the effectiveness of treatment methods. Parallels can be drawn with
the US Environment Protection Agency (EPA) standards, which differentiate
between Class A and Class B sludge products.
The current Code of Practice and Regulations will be amended by DEFRA to take
account of the ‘Safe Sludge Matrix’.
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Following discussions between the major stakeholders in 2000, agreement was
reached to allow untreated sewage sludge to continue to be applied to land used to
grow specified industrial crops that do not normally have a food use, until 31
December 2005. After that date untreated sewage sludge can no longer be applied
to agricultural land (ADAS 2001b). The specified crops include willow and
poplar, for coppicing; Miscanthus, for biomass; hemp, for fibre; and high erucic
acid rape (HEAR), for the production of high erucic acid rape oil (HERO).
Impact and implications for bioremediation
The Safe Sludge Matrix provides a framework of guidance within which all
operators work when applying sewage to agricultural land. The Matrix is not
likely to limit or hinder the use of wastes on bioenergy crops for bioremediation.
The agreement will, however, ban the use of untreated sewage sludge to all crops
from 31 December 2005. The ‘Industrial Crops Agreement’ permits the use of
untreated sewage sludge, including sludges that fail the end product standards for
pathogens, on crops such as Miscanthus and short rotation coppice until 31
December 2005. This may actually be helpful in stimulating interest in these
crops, as sewerage operators look for secure outlets in the wake of the foot and
mouth disease crisis.
5.2.10
Sewage sludge use in forestry and on restored land
The use of sewage sludge on non-agricultural land, including forestry, and for land
restoration is outside the scope of the Sludge (Use in Agriculture) Regulations (as
amended) (SI, 1989 & 1990)and the Code of Practice for Agricultural Use of
Sewage Sludge (DoE, 1996). These activities are however controlled by the Waste
Management Licensing Regulations 1994 (SI, 1994b).
Applications are exempt from the waste licensing, provided that the sludge
provides ecological benefit and does not exceed the metal limits as set out in
Schedule 2 of the Sludge (Use in Agriculture) Regulations (SI, 1990). It is,
however, necessary to register the activity with the Environment Agency.
Further guidance can be found in a Forestry Commission Bulletin – A Manual of
Good Practice for the Use of Sewage Sludge in Forestry (Forestry Commission,
1992).
Impact and implications for bioremediation
The regulations and information bulletin are discussed in other sections and
provide a framework of guidance within which all operators work when applying
sewage to non-agricultural land. They are not likely to limit or hinder the use of
wastes on bioenergy crops for bioremediation. The Industrial Crops Agreement
will however affect the use of sludge on short rotation coppice, but is not thought
to limit uptake in any way.
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5.2.11
Waste Management Licensing Regulations
The Waste Management Licensing Regulations (SI, 1994) deal with all aspects of
the storage, treatment, and recovery or disposal of controlled wastes, including the
licensing of facilities and operators. Wastes arising from agriculture, as defined by
the 1947 Agricultural Act (HMSO, 1947), are excluded from this Regulation.
However, under these regulations, certain controlled wastes for use in agriculture
are exempt from the requirements of the licensing regulations. These wastes are
listed in Schedule 3 of the Regulations (Table 2 of paragraph 7), and include waste
soil or compost; waste wood, bark or other plant material; waste food, drink or
materials used in or resulting from the preparation of food or drink; blood and gut
contents from abattoirs; waste lime; lime sludge from cement manufacture or gas
processing; waste gypsum; paper waste sludge, waste paper and de-inked paper
pulp; dredgings from inland waterways; textile waste; sludge from biological
treatment plants; and waste hair and effluent treatment sludge from a tannery.
Such exempt activities must however result in benefit to agriculture or ecological
improvement and are subject to an annual maximum application of 250 tonnes ha-1,
or in the case of dredgings from inland waters, 5,000 tonnes ha-1, and are subject to
pre-notification to the regulator.
The regulations also require that waste is recovered or disposed of without
endangering human health and without using processes or methods which could
harm the environment and in particular without:
 risk to water, air, soil, plants or animals; or
 causing nuisance through noise or odours; or
 adversely affecting the countryside or places of special interest.
The Regulations are due to be revised in the near future.
Impact and implications for bioremediation
The regulations provide a framework of guidance within which all operators work
when applying controlled wastes to agricultural land. They are not likely to limit
or hinder the use of appropriate wastes on bioenergy crops for bioremediation
although, due to the definitions used in the regulations, there can be difficulties in
categorising certain wastes. The revised regulations will, however, ban the use of
septic tank wastes and are likely to require more pre-notification and auditing of
applications.
The meanings of the terms “benefit to agriculture” and “ecological improvement”
are not clearly defined in the regulations and are open to interpretation by waste
producers and their contractors, and by the Environment Agency. As a result, a
waste can sometimes be accepted in one region and prohibited in another. It is
hoped that clearer definitions, and an indication of what evidence is required to
demonstrate ‘benefit’, will be included in future guidance.
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5.3
UK LEGISLATION, NOT PRIMARILY AIMED AT WASTE
DISPOSAL, WHICH MAY AFFECT ORGANIC WASTES
5.3.1 Animal By-Products Order
The Animal By-Products Order 1999 (ABPO) (SI, 1999), amending earlier orders
in 1992 and 1996, controls all material derived from animals which is not intended
for human consumption. It applies to any unprocessed animal by-products (animal
carcasses and parts of animal carcasses which are not intended for human
consumption) and to catering waste which contains meat or other products of
animal origin. It does not apply to BSE suspects, ‘specified risk material’ (SRM),
animals suspected of carrying a notifiable disease or stock slaughtered under the
‘over thirty month’ scheme.
The ABPO requires that such animal by-products be consigned to rendering,
incineration, or other permitted routes. It permits landfilling and/or treatment of
catering waste, but requires that ruminant animals, pigs and poultry do not gain
access to the waste. Catering waste, by definition, appears to include domestic
waste and, therefore, would also include non-source segregated municipal solid
waste (MSW).
Guidance notes indicate that ‘poultry’ includes wild birds and that composting and
biodigestion are not permitted disposal routes for animal by-products under the
ABPO.
Impact and implications for bioremediation
The regulations were designed to control the spread of animal diseases and are
necessary, particularly given recent outbreaks of swine fever and foot and mouth
disease. However, current interpretation of the regulations effectively prevents the
use of open air composting of MSW - since it is virtually impossible to prevent
access by wild birds. Furthermore, recent guidance from DEFRA would suggest
that, even after treatment, access by wild birds would need to be prevented, thereby
effectively preventing the recycling of any MSW to agricultural or restored land.
If this interpretation were correct then the ABPO would limit the use of such
wastes on bioenergy crops.
The UK ABPO regulations are due to be superceded by new EU regulations
currently being agreed by the Commission and it is understood that land
application of certain, lower risk wastes will be permitted as long as they undergo
‘pre-treatment’ to specified conditions of temperature and pressure.
The Environment Agency are due to publish the results of a risk assessment of
such wastes and it is understood that the findings will be used as the basis of any
new guidance on the treatment of wastes that may contain animal by-products.
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5.3.2 Plant Health (Great Britain) Order 1993
Under Article 20 of the Plant Health (Great Britain) Order 1993 (as amended) (SI,
1996) growers are required to notify Agricultural Departments of the known or
suspected presence of any quarantine pest or harmful organism not normally found
in Great Britain.
Guidance on minimising the risks of transmitting organisms or disease through the
recycling or land application of waste material, including soil or growing media,
wash water, trimmings and peelings and outgrades or other plant material can be
found in PB3580 (MAFF/WOAD/SOAEFD, 1998).
Rhizomania and potato brown rot, as well as more common pests and diseases such
as Phytophthora, Fusarium, Verticillium and stem and bulb nematodes, can all be
potentially spread through waste products and care should be taken to ensure
appropriate treatment of wastes is undertaken.
Impact and implications for bioremediation
The Code of Practice, and associated regulations, provide a framework of guidance
within which operators work when applying potentially infected wastes to
agricultural land. They are not likely to limit or hinder the use of appropriate
wastes on bioenergy crops for bioremediation.
5.3.3 Planning controls
The Town and Country Planning Act 1990 (as amended by the Planning and
Compensation Act 1991) controls the use of land for mineral excavation.
Assuming consent is granted, detailed planning conditions will be established
which control the working, restoration and aftercare of such sites. The
Environment Act 1995 provides for a review and updating of permissions granted
between 1950 and 1980, with periodic reviews afterwards.
Impact and implications for bioremediation
The regulations provide a framework of guidance within which all operators work.
The planning conditions are often very detailed and may define the type of
products that should be used in the restoration process. This may limit the use of
alternative wastes or products that can be used. Detailed conditions will only
apply to more recent consents; older consents may have little or no aftercare
requirements. Many derelict former industrial sites fall into this latter category.
5.3.4 Contaminated land
Contaminated land may be a statutory nuisance under the Environmental
Protection Act 1990 (HMSO, 1990) and the local authority may have the power, in
some circumstances, to issue a notice requiring the person responsible, or the
owner or occupier, to prevent the nuisance.
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Under the Environment Act 1995 (HMSO, 1995) local authorities are required to
identify contaminated land which poses a significant risk of harm to human health
or the environment generally. The owner or person responsible may then be
required to remedy the problem. Local authorities may also, under certain
circumstances, fund the clean-up of contaminated land.
Impact and implications for bioremediation
The regulations provide a framework of guidance and may well be a driver for
identifying sites, which require bioremediation.
5.4
EC LEGISLATION: PROPOSALS THAT MAY AFFECT ORGANIC
WASTES
5.4.1 Introduction
A range of measures are currently proposed at an EU level that may also impact on
the use of organic wastes for bioremediation, as the proposals - if adopted – will, in
time, have to be adopted into UK legislation.
5.4.2 EC Sludge Directive – Working Document 3rd Draft
Sewage sludge applications to land are currently regulated by the EC Sludge
Directive (EC, 1986). During 1999 and 2000, a series of national expert meetings
were held, with a view to revising the Sludge Directive. The Commission has
issued its 3rd Working draft although it is understood that this may not represent the
final proposals. Revised proposals, which will be subject to consultation and
formal agreement with Commissioners and Member States, are expected in spring
2003.
The draft proposals of the EC Sludge Directive include:
 significant reductions in the maximum levels for heavy metals in soils (this
would cause major difficulties in the UK)
 inclusion of septic tank, food wastes, paper wastes and other manufacturing
wastes within the remit of the directive
 inclusion of treatment standards, based on the degree of pathogen reduction
achieved (similar to the proposed UK regulations)
 introduction of maximum levels of heavy metals in the sludge itself, and
 introduction of limits on organic compounds (based on the limits in Annex
III, all UK sewage sludges would fail the limits for PAH).
The EU are keen to progress with the concept of sustainability, based on the idea of
zero addition of contaminants; whereas the UK is proposing limits based on
scientific studies and a precautionary principle approach. There is also concern
about the absence of agreed analytical methods and protocols.
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Impact and implications for bioremediation
The Directive is still far from agreed and it is likely that further compromises will
be made before it is finalised. In any event it is not likely to be ratified until 2003
and is, therefore, unlikely to be enacted in UK legislation until about 2005. It
would, however, severely restrict the recycling of sewage sludges to agricultural
land and may lead to problems with other ‘wastes’ that will fall within the scope of
the directive - since even less is known about the level of organic contaminants, for
example, that may be present in other sludges. It also confirms the principle that if
you mix any sludge with other non-regulated wastes, then the entire combined
waste stream becomes affected by the Directive. This potentially limits some
treatment sites where sewage sludge is either used as a feedstock or is treated in
order to receive a gate fee, to subsidise the treatment of other wastes.
5.4.3 EC Biological Treatment of Biowaste Directive – Working
Document 2nd Draft
Often referred to as the ‘Compost Directive’ this new directive is still being
discussed and debated by the various national experts. The current working
document is intended as a basis for preliminary discussions. Its objectives are to
promote the treatment of bio-waste, and to reduce any negative impact of such
wastes. It also seeks to protect the soil and ensure human and animal health is not
affected by the use of treated or untreated bio-waste.
The draft promotes the waste hierarchy:
 Prevent or reduce production
 Re-use
 Recycle
 Compost or anaerobically digest and use in agriculture
 Mechanical/biological treatment of bio-waste
 Use bio-wastes as a source for generating energy
The document promotes home composting; on-site composting and anaerobic
digestion; ‘community composting’; separate collection schemes (targets are
proposed) and promotes the mechanical/biological treatment of any residual
municipal waste.
The document also proposes the introduction of end-product standards and
suggests two classes of compost/digestate, as well as figures for stabilised biowaste. There appears to be a reasonable ‘read-across’ to the sludge directive for
organic contaminant limit values – but, based on a quick review of the figures for
heavy metals, it would appear that the proposed levels may be difficult to achieve
from mixed waste streams and appear to have been based on green waste
composting.
There are also some concerns about the choice of indicator organisms i.e.
Salmonella senftenberg and Clostridium perfringens. Although both are under
review, the UK has historically used E. coli and Salmonella spp. as indicator
organisms.
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The draft also reinforces the need for any reuse on land to result in agricultural
benefit or ecological improvement. It also proposes the following limits on
application rates:
 Class 1 – no limit, but used according to best agronomic practice.
 Class 2 – 30 t dry matter ha-1 (three year average)
 Stabilised bio-waste – no repeat application in 10 years and a limit of 200 t
dry matter ha-1 in a single application.
All applications must be made in accordance with the record keeping and soil
sampling requirements of the Sludge Directive (EC, 1986).
Impact and implications for bioremediation
The Directive is still far from agreed and it is likely that further amendments will
be made, in light of negotiation over the Sludge Directive. It would, however, if
implemented in the draft form, severely restrict the treatment of many mixed waste
streams and industrial wastes. It also confirms the principle that if you mix any
digestate with other non-regulated materials, such as peat or inorganic fertilisers,
then all of the combined material becomes affected by the Directive. This
potentially limits some treatment sites where wastes are either used as a feedstock
or are treated in order to receive a gate fee, to subsidise the treatment of other
wastes, and also limits the opportunity to add value. As they currently stand,
however, the proposed application rate limits would severely restrict the use of biowastes in land reclamation and bioremediation since current practice is to apply
much higher rates than are proposed in the second working draft.
5.4.4 EC Landfill Directive (EC 1993/31)
The recently adopted Landfill Directive 1993/31/EC (EC, 1993) requires that
significant reductions must be made in the amount of putrescibles being disposed
of to landfill sites by 2006, further targets also apply for 2009 and 2016.
Government estimates that 3.2 m tonnes per year must be diverted from landfill to
meet initial targets. While the paper ‘Limiting Landfill’ (DEFRA, 1999)
recognises that a significant part will be played by waste minimisation initiatives, it
remains focused on treatment of waste as an alternative to disposal to landfill.
Surveys of household waste (MSW) identify 20% as organic food, and 33%
paper/card, giving a potential 53% of MSW material that can be composted. It is
estimated that 5 million tonnes per annum of organic material is disposed of within
the MSW stream, which could be realistically diverted if a suitable product and
market could be found.
The Directive is driving waste management companies and waste producers to
look at land spreading and bioremediation as a possible route to divert material
away from landfill.
The Directive was implemented in England and Wales under the Pollution
Prevention and Control (England and Wales) Regulations 2000, as amended (SI,
2002).
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Impact and implications for bioremediation
This directive is acting as a driver for change and is encouraging the development
of alternative routes for recycling, either in the form of waste treatment or the
establishment of new markets for recycling of wastes. Recent concern from
retailers and food processors has led many waste management operators to look at
the use of organic wastes for bioremediation and biomass production, since they
are seen as more ‘secure, than traditional outlets.
5.5
REFERENCES
ADAS (2000). MANNER User Guide. ADAS, ADAS Gleadthorpe Research
Centre, Mansfield, Nottingham.
ADAS (2001a). The Safe Sludge Matrix – Guidelines for the Application of
Sewage Sludge to Agricultural Land. ADAS, Wolverhampton.
ADAS (2001b). Guidelines for the Application of Sewage Sludge to Industrial
Crops. ADAS, Wolverhampton.
DEFRA (1999). Limiting Landfill: A consultation paper on limiting landfill to
meet the EC Landfill Directive’s targets for the landfill of biodegradable
municipal waste. Originally published by DETR. http://www.defra.gov.uk/
environment/waste/strategy/landfill/index.htm.
DETR/MAFF/WO (1997). Draft regulations establishing the action programme
measures to apply in Nitrate Vulnerable Zones in England and Wales Consultation Document. EC Nitrate Directive (91/676/EEC): Department of
the Environment, Transport and the Regions, Ministry of Agriculture,
Fisheries and Food, Welsh Office. December 1997.
DoE (1990).
Environmental Protection Act, 1990.
Environment. HMSO, London.
Department of the
DoE (1993). Sludge Use in Agriculture 1990/91. UK report to the EC
Commission under Directive 86/278/EEC. Department of the Environment,
London.
DoE (1996). Code of Practice for the Agricultural Use of Sewage Sludge.
Department of Environment. HMSO, London.
EA (2000). Aquatic Eutrophication in England and Wales – a management
strategy. Environment Agency.
EA (2001). Water pollution incidents in England and Wales 1999. Environment
Agency.
EC (1991a). Council Directive on Waste (91/156/EEC) of 18 March 1991
amending Directive 75/442/EEC on waste. Official journal of the European
Community L 078, 26/03/1991. pp 0032–0037.
EC (1991b). Nitrate Directive (91/676/EEC). Official Journal of the European
Community.
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EC (1991c). Urban Waste Water Treatment Directive (91/271/EEC). Official
Journal of the European Community, L13530.5.91, p 40. As amended by
Council Directive 98/15/EC, February 1998.
Summary of Measures
Implemented by Member States COM (1998).
EC (1993). Landfill Directive (1993/31/EEC). Official Journal of the European
Community.
Forestry Commission (1992). A Manual of Good Practice for the Use of Sewage
Sludge in Forestry. Bulletin 107. Forestry Commission Publications,
Wetherby, West Yorkshire.
HMSO (1947). Agriculture Act 1947. Chapter 48. HMSO, London.
HMSO (1990a).
London.
Environmental Protection Act 1990.
Chapter 43.
HMSO,
HMSO (1990b). Town and Country Planning Act 1990. Chapter 8. HMSO,
London.
HMSO (1991). Water Resources Act 1991. Chapter 57. HMSO, London.
HMSO (1995). Environment Act 1995. Chapter 25. HMSO, London.
MAFF (1998a). The Water Code. Code of Good Agricultural Practice for the
Protection of Water. Ministry of Agriculture, Fisheries and Food/Welsh
Office Agriculture Department. MAFF Publications, London. (PB0585).
MAFF (1998b). The Air Code. Code of Good Agricultural Practice for the
Protection of Air. Ministry of Agriculture, Fisheries and Food/Welsh Office
Agriculture Department. MAFF Publications, London (PB0618). 74 pp.
MAFF (1998c). The Soil Code. Code of Good Agricultural Practice for the
Protection of Soil. Ministry of Agriculture, Fisheries and Food/Welsh Office
Agriculture Department. MAFF Publications, London. (PB0617).
MAFF (2000). Fertiliser Recommendations for Agricultural and Horticultural
Crops. 7th edition. MAFF Reference Booklet 209. HMSO, London.
MAFF/WOAD (1998). Guidelines for Farmers in NVZs. MAFF Publications,
London (PB3277).
MAFF/WOAD/SOAEFD (1998). Code of Practice for the Management of
Agricultural and Horticultural Waste. Ministry of Agriculture, Fisheries and
Food/Welsh Office Agriculture Department/Scottish Office Agriculture
Environment and Forestry Department.
MAFF Publications, London.
(PB3580).
SI (1989). United Kingdom Statutory Instrument No. 1263. The Sludge (Use in
Agriculture Regulations, 1989. HMSO, London.
SI (1990). United Kingdom Statutory Instrument No. 880. The Sludge (Use in
Agriculture) (Amendment) Regulations, 1990. HMSO, London.
SI (1991). United Kingdom Statutory Instrument No. 324. The Control of
Pollution (Silage, Slurry and Agricultural Fuel Oil) Regulations 1991.
HMSO, London.
SI (1992). United Kingdom Statutory Instrument No. 3293. The Animal Health
Act 1981 (Amendments) Regulations 1992. HMSO, London.
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SI (1994a). Statutory Instrument No. 2841. Urban Waste Water Treatment
(England & Wales) Regulations 1994. HMSO, London.
SI (1994b). Statutory Instrument No. 1056. The Waste Management Licensing
Regulations 1994. HMSO, London.
SI (1996). United Kingdom Statutory Instrument No. 3242. Plant Health (Great
Britain) (Amended) (No 3) Order 1996. HMSO, London.
SI (1998). Statutory Instrument No. 1202. The Action Programme for Nitrate
Vulnerable Zones (England and Wales), 1998. HMSO, London.
SI (1999). Statutory Instrument No. 646. The Animal By-Products Order 1999.
HMSO, London.
SI (2002). United Kingdom Statutory Instrument No. 275. The Prevention and
Control (England and Wales) (Amendment) Regulations 2002. HMSO,
London.
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CHAPTER 6
CONCLUSIONS AND RESEARCH RECOMMENDATIONS
6.1
CONCLUSIONS
The planned expansion of the area in the UK planted with energy crops, to partially
facilitate the requirement that 10% of all electricity generation in 2010 must come
from renewable sources, provides an opportunity for the safe disposal of organic
wastes and for the bioremediation of polluted, ex-industrial land.
The safe disposal of organic wastes, such as sewage sludge, animal manures and
slurries, paper sludge and abattoir wastes can be problematic. Many of these
materials contain heavy metals, organic pollutants or pathogens that are potentially
damaging to human health; and their utilisation on food crops causes public
concern, despite stringent safeguards embodied in various articles of legislation
and codes of practice.
The opportunity to apply such wastes to vigorous non-food crops, with extensive
and rapidly developing root systems, would appear to have obvious attractions –
offering an environmentally-friendly alternative to food crop applications, landfill
or incineration. The facts that energy crops are regularly harvested, with the
above-ground parts being removed from the site and burnt, and that the postharvest period (in late winter and early spring) offers ready access to the crop for
waste spreading, would appear to further increase the ‘bioremediation’ potential.
One further, but very important, benefit that might accrue to landowners exploiting
the waste utilisation or bioremediation potential of energy crops is improved
financial viability. Current levels of government grants and anticipated income
from crop sales are proving insufficiently attractive to entice more than a small
minority of landowners into energy crop production. Any opportunity to ‘add
value’ to energy crops, such as providing a regular disposal point or organic
wastes, will increase their attractiveness to many potential growers.
This review has confirmed many of the potential opportunities for waste
utilisation/bioremediation that energy crops offer. However, it has also highlighted
certain environmental risks that need to be considered. For example…

Repeated applications of organic wastes with a high P content (e.g. pig or
poultry manures), at rates which apply the maximum permitted quantity of
N (i.e. 250 kg N ha-1 yr-1), are likely to supply P in excess of the maximum
permitted under the relevant code of practice. This will increase the risks
of eutrophication of water-bodies.

Similarly, repeated applications of organic manures (livestock manure,
biosolids, industrial waste) will be likely to add heavy metals in excess of
crop uptake, leading to an accumulation in the soil. Heavy metals can also
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be carried through to the ash left after burning, so disposal of ash must be
made with due regard to its chemical composition.

Slurry applications might lead to excessive ammonia volatilisation, and
local odour problems – as the most effective way of reducing volatilisation
is to inject the slurry into the soil, but this is not an option in energy crops.

Very little is known about the ecological impacts of waste applications to
energy crops. From the limited amount of information that is available, it
can be surmised that the generally low conservation value of actively
managed energy crops is unlikely to be significantly reduced or increased.
However, the potential bioaccumulation of heavy metals in predator species
would be an issue of some concern – if large areas of energy crops around,
for example, a single power station were to be regularly treated with wastes
containing high metal concentrations.
The genetic modification of biomass crops to facilitate increased pollutant uptake,
transport, accumulation and tolerance offers the potential to dramatically increase
the effectiveness of phytoremediation of organic compounds and metals from
contaminated sites. Although the genetics of poplars have been extensively
studied, very little research has so far been undertaken on the genetic modification
of the primary candidates for energy cropping in the UK - willows and Miscanthus.
Consequently, the realisation of this potential is still some distance away.
6.2
RESEARCH RECOMMENDATIONS
The following subjects should be considered as priority areas for research:
1.
Nutrient and CO2 losses to the environment on removal of biomass crops and
return of land to conventional agriculture.
2.
The nutrient requirements of energy crops under a range of conditions.
3.
Investigation of the potential use of leaf analysis in the assessment of nitrogen
requirements for energy crops.
4.
Further research on species or varietal sensitivity of SRC and Miscanthus
crops to landfill leachate and the maximum conductivity (ec) of leachates for
‘safe’ irrigation.
5.
N2O efflux from sites without organic amendments, but with and without
inorganic nutrient fertilisation, and in the presence of normal litter quantities,
to establish a baseline flux balance.
6.
N2O and CH4 flux after organic amendment application in normal SRC
conditions i.e. after the passage of harvest machinery and in the presence of
litter and stools.
7.
Contribution of both applied organic wastes and annual litter fall to medium
and long-term soil organic matter carbon sequestration.
8.
Longevity and distribution of carbon sequestered in woody root systems of
SRC, and the turnover of root biomass in Miscanthus.
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9.
Longer-term N2O and CH4 flux from Miscanthus sites receiving annual
applications of organic/inorganic fertilisers.
10. Potential of the mineralisation efflux of CO2 from frequently applied organic
wastes to result in carbon fertilisation of C3 plants such as willow and poplar,
by a CO2-enriched micro-climate.
11. Further screening of willow, poplar and energy grass varieties to evaluate their
levels of tolerance to inorganic and organic pollutants, and assessment of the
extent to which pollutants are either inactivated (e.g. certain organic
compounds) or accumulated in harvestable plant parts.
12. Assessment of the types and quantities of organic contaminants in organic
manures and rates of uptake and breakdown by energy crops.
13. The genetic modification of energy crops to increase their bioremediation
potential.
14. The agronomic and ecological impacts of applying different waste materials
(at different rates) to energy crops.
15. Environmental fate of pollutants in contaminated soils after establishment of
energy crops.
16. Environmental fate of pollutants in waste materials after their application to
energy crops.
17. The nature and environmental fate of pollutants in flue emissions and ash after
combustion/pyrolysis for power generation.
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