DEFRA WQ0206: AGRONOMIC BENEFITS AND …sciencesearch.defra.gov.uk/Document.aspx?Document=DEFRA8021App… · sample that can be combusted at 550oC. The loss in weight during combustion

DEFRA

WQ0206: AGRONOMIC BENEFITS AND ENVIRONMENTAL IMPACTS OF SPREADING ORGANIC MATERIALS TO LAND

APPENDIX A: ASSESSMENT OF THE POLLUTION RISK, GREENHOUSE GAS EMISSIONS AND AGRONOMIC BENEFITS OF ORGANIC MATERIALS RECYCLED TO LAND

WRc Ref: DEFRA8021.02

JANUARY 2010

CONTENTS

GLOSSARY OF TERMS AND CONCEPTS 1

A1. INTRODUCTION 3

A1.1 Overview 3 A1.2 Objectives 3 A1.3 Background 4 A1.4 Approach 5 A1.5 Appendix Structure 5

A2. DATA COLLECTION AND RELEVANT WASTE CHARACTERISTICS 7

A2.1 Waste categories 7 A2.2 Waste characteristics considered 10 A2.3 Waste characteristics background 11 A2.4 Emissions to air 22 A2.5 Characteristics by waste type 26

A3. METHODOLOGY FOR ASSESSMENT OF AGRONOMIC BENEFIT 30

A3.1 Overview 30 A3.2 Calculation of loadings by normalisation to N fertiliser value 30 A3.3 Calculation of loadings of wastes as soil conditioner 32 A3.4 Comparison of ammonia emissions from loadings 33 A3.5 Nitrous oxide emission estimation 35 A3.6 Carbon dioxide emission 36 A3.7 Methane emissions 36 A3.8 Phosphorus nutrient value assessment 36 A3.9 K, Mg and S nutrient value assessment 36 A3.10 Assessment of metal applications 37 A3.11 Assessment of organic pollutants 37 A3.12 Assessment of pathogens 37

A4. ASSESSMENT AGRONOMIC VALUE OF WASTES AS FERTILISER 38

A4.1 Total dry matter loading comparison 38 A4.2 Total nitrogen loadings 40 A4.3 Phosphorus loadings 41 A4.4 Loadings of other major nutrient 43 A4.5 Metal loadings from application of wastes to a PAN of 100 kg N/ha 49

A5. ASSESSMENT AGRONOMIC VALUE OF WASTES APPLIED AS SOIL CONDITIONERS 51

A5.1 Total dry matter loadings for different wastes 51 A5.2 Total N applications when applying wastes to increase SOC 52 A5.3 Loadings of PAN for applications to increase SOC 54 A5.4 Phosphorus loadings for applications to increase SOC 55 A5.5 Loadings of other major nutrients K, Mg and S 56 A5.6 Metal loadings 61

A6. GASEOUS EMISSIONS TO AIR AND CARBON SEQUESTRATION 62

A6.1 Introduction 62 A6.2 Emissions of ammonia to air 62 A6.3 Carbon dioxide emissions and carbon sequestration 67

A7. ORGANIC POLLUTANTS 72

A7.1 PAH and PCB assessment for waste applied as soil conditioners. 73

A8. GROUND AND SURFACE WATER RISK ASSESSMENT 74

A8.1 Introduction 74 A8.2 Methodology 74 A8.3 Concentration input 77 A8.4 Comparison of treatment method for sewage sludge 89 A8.5 Conclusion 103

A9. PATHOGENS 104

A9.1 Introduction 104 A9.2 Animal by-products 104 A9.3 Pathogens in compost 106 A9.4 Pathogens in sewage sludge 107 A9.5 Pathogens in manures 108 A9.6 Pathogens in paper sludges 109

A10. KEY FINDINGS AND CONCLUSIONS 110

A10.1 Comments on general approach 110 A10.2 Agronomic benefit 110 A10.3 Soil quality issues 111 A10.4 Greenhouse gas emissions 112 A10.5 Pathogen risks 113 A10.6 Suggested waste characterisation required 114

REFERENCES 115

A11. APPENDIX 127

A11.1 Characteristics of waste types used in this assessment (mean values) 127 A11.2 Characteristics of waste treatments used in this assessment (mean values) 128 A11.3 Heavy metal loadings from waste types applied to give a PAN of 100 kg

N/ha 129 A11.4 Heavy metal loadings from waste treatments applied to give a PAN of 100

kg N/ha 130 A11.5 Heavy metals loadings from waste types applied as soil conditioner 131 A11.6 Heavy metal loadings from waste treatments applied as soil conditioner 132

LIST OF TABLES

Table 1-1 Typical application rate of inorganic fertiliser to soils (kg/ha) 5

Table 2-1 Generic organic wastes types applied to soil 7

Table 2-2 Waste types and colours used in this study 8

Table 2-3 Waste treatment codes used in this study 10

Table 2-4 Percentage carbon mineralised in soils after 70 days incubation for wastes composted for different times 18

Table 3-1 Estimation of organic nitrogen mineralisation factor from waste C/ON ratio 31

Table 3-2 Characteristics of mature compost for example PAN calculation 32

Table 3-3 Example estimation of ammonia emission factor (EF) as percentage of total ammoniacal nitrogen (TAN) 35

Table 4-1 Comparison nutrient content in sewage sludge and composts (g/kg DM) 48

Table 4-2 Summary of waste metal application rates that are close to or exceed sewage sludge application limits (wastes applied to PAN of 100 kg/ha) 49

Table 5-1 Summary of wastes and waste treatments with high nutrient loadings when applied as soil conditioner 60

Table 7-1 Example analysis of impact of selected wastes applied as soil conditioner on soil organic pollutant levels 73

Table 8-1 Statutory benchmarks used for risk assessment 74

Table 8-2 Soil characteristics used for risk assessment 76

Table 8-3 Summary table of 95%ile concentrations (mg/kg DM) used in the groundwater and surface water risk assessment 78

Table 8-4 Summary table of 95%ile leachate concentrations (mg/l) calculated by Consim for the groundwater and surface water risk assessment with benchmark used. 87

Table 8-5 Table comparing highest 95%ile leachate concentration with surface and groundwater threshold concentrations (EQSs and DWLs). 89

Table 10-1 Suggested waste characterisation testing required to improve prediction of agronomic benefit and environmental impact of organic wastes spread to land 114

LIST OF FIGURES

Figure 4.1 Dry matter loadings (t/ha) for different wastes to achieve PAN of 100 kg N/ha) (mauve bar - sewage sludge standard where PAN is 50% TN) 38

Figure 4.2 Loadings of waste treatments to achieve the PAN loading of 100 kg N/ha compared with benchmark sewage sludge (purple bar) 39

Figure 4.3 Total N loadings for waste types applied to achieve a PAN of 100 kg N/ha 40

Figure 4.4 Total N loadings for waste treatments applied to achieve PAN of 100 kg N/ha 41

Figure 4.5 Available P loadings from application of waste types to a PAN of 100 kg N/ha 42

Figure 4.6 Available P loadings for waste treatments applied to a PAN of 100 kg N/ha 43

Figure 4.7 Available K loadings for different wastes types to give PAN of 100 kg N/ 44

Figure 4.8 Available K loadings for different waste treatments applied to give a PAN of 100 kg N/ha 44

Figure 4.9 Total Mg loadings for different waste types applied to a PAN of 100 kg N/ha 45

Figure 4.10 Total Mg loadings for different waste treatments applied to PAN of 100 kg N/ha 46

Figure 4.11 Total S loadings for waste types applied to achieve PAN of 100 kg N/ha 47

Figure 4.12 Total S loadings for waste treatments applied to PAN of 100 kg N/ha 48

Figure 5.1 Loadings of waste types as soil conditioner 51

Figure 5.2 Loadings of waste treatments as soil conditioner 52

Figure 5.3 Total N loadings for waste types applied as soil conditioner 53

Figure 5.4 Total N loadings for waste treatments applied as soil conditioner 54

Figure 5.5 Loadings of PAN for waste types applied as soil conditioner 54

Figure 5.6 Loadings of PAN for waste treatments applied as soil conditioner 55

Figure 5.7 Available P loadings from waste types applied as soil conditioner 56

Figure 5.8 Available P loading from waste treatments applied as soil conditioner 56

Figure 5.9 Available K for waste types applied as soil conditioner 57

Figure 5.10 Available K for waste treatments applied as soil conditioner 58

Figure 5.11 Total Mg loadings for waste types applied as soil conditioner 58

Figure 5.12 Total Mg loadings for waste treatments applied as soil conditioner 59

Figure 5.13 Total S loadings for waste types applied as soil conditioner 59

Figure 5.14 Total S loadings for waste treatments applied as soil conditioner 60

Figure 6.1 Ammonia emissions from application of waste types to PAN of 100 kg N/ha 63

Figure 6.2 Ammonia emissions from waste types applied as soil conditioner 63

Figure 6.3 Nitrous oxide emissions from waste types applied to PAN of 100 kg N/ha 64

Figure 6.4 Nitrous oxide emission from waste treatment applied to PAN of 100 kg N/ha 65

Figure 6.5 Nitrous oxide emissions from waste types applied as soil conditioner 66

Figure 6.6 Nitrous oxide emissions from waste treatments applied as soil conditioner 66

Figure 6.7 CO2 emissions from waste types applied to PAN of 100 kg N/ha 67

Figure 6.8 CO2 emissions from waste treatments applied to PAN of 100 kg N/ha 68

Figure 6.9 CO2 emissions of waste types applied as soil conditioner 69

Figure 6.10 CO2 emissions from waste treatments applied as soil conditioner 69

Figure 6.11 Estimated C sequested in soil from application of waste types to PAN of 100 kg n/HA 70

Figure 6.12 Estimated C sequested in soil from application of waste treatments to PAN of 100 kg N/ha 71

Figure 8.1 Total nitrogen concentration in different waste types 79

Figure 8.2 Extractable nitrogen concentration in different waste types 80

Figure 8.3 Total phosphorus concentration in different waste types 80

Figure 8.4 Available phosphorus concentration in different waste types 81

Figure 8.5 Total potassium concentration in different waste types 81

Figure 8.6 Available potassium concentration in different waste types 82

Figure 8.7 Zinc concentration in different waste types 82

Figure 8.8 Copper concentration in different waste types 83

Figure 8.9 Nickel concentration in different waste types 83

Figure 8.10 Lead concentration in different waste types 84

Figure 8.11 Cadmium concentration in different waste types 84

Figure 8.12 Chromium concentration in different waste types 85

Figure 8.13 Mercury concentration in different waste types 85

Figure 8.14 Total nitrogen concentration by treatment type 91

Figure 8.15 Total phosphorus concentration by treatment type 92

Figure 8.16 Nickel concentration by treatment type 93

Figure 8.17 Lead concentration by treatment type 94

Figure 8.18 Cadmium concentration by treatment type 95

Figure 8.19 Chromium concentration by treatment type 96

Figure 8.20 Mercury concentration by treatment type 97

Figure 8.21 Zinc concentration by treatment type 98

Figure 8.22 Copper concentration by treatment type 99

Figure 8.23 Comparison of WW and DM for total nitrogen concentration in sewage sludge waste 100

Figure 8.24 Comparison of WW and DM for total phosphorus concentration in sewage sludge waste 101

Figure 8.25 Comparison of WW and DM for lead concentration in sewage sludge waste 102

GLOSSARY OF TERMS AND CONCEPTS

Term Description

AD Anaerobic digestion

Biodegradability Biodegradability of a waste is a measure of the degree to which the plant or animal derived organic matter is decomposed by microbes during biological treatments such as aerobic composting, anaerobic digestion or following disposal in a landfill.

CAN Crop available nitrogen

DM Dry matter. Mass of material after complete drying (to constant weight) at 105oC. Expressed as % wet weight.

DOC Dissolved organic carbon

GHG Greenhouse gases – Gases that contribute to the greenhouse effect and climate change: water vapour, nitrous oxide, methane, carbon dioxide and ozone.

LOI Loss on Ignition. A measure of the quantity of organic matter in the sample that can be combusted at 550oC. The loss in weight during combustion equates to the mass of organic matter in the sample. It is expressed in % of DM.

EQS Environmental quality standards

DWS drinking water standards

CLO Compost-like outputs

SOC Soil organic carbon

SNS Soil nitrogen supply

SGV Soil guidline value

MBT Mechanical and biological treatment

MSW Municipal solid waste

BMW Biodegradable municipal waste

Cd Cadmium

Pb Lead

Hg Mercury

Cu Copper

Cr Cromium

Zn Zinc

As Arsenic

ON Organic nitrogen – Nitrogen present in nitrogenous organic compounds such as proteins and amino acids. This form of nitrogen is not available for immediate uptake by the crop and must be first broken down to mineral nitrogen.

PAN Potentially available nitrogen

RAN Readily available nitrogen – Nitrogen present in mineral form (NH4+ or

nitrate), available for immediate uptake by the crop

RB209 MAFF: Fertiliser recommendations for agricultural and horticultural crops (RB209): Seventh edition (2000)

RSD (%) Relative Standard Deviation is the positive square root of the variability of a data set, as a percentage of the mean of the group of data

TAN Total ammonia nitrogen

TN Total nitrogen

WRc Ref: DEFRA8021.01 January 2010

3

A1. INTRODUCTION

A1.1 Overview

There has been substantial research on the benefits and impacts of spreading organic materials to agricultural land. For some materials, such as sewage sludge and animal manures, characterisation data exists pertinent to the benefits and environmental impact of this material when it is recycled to land. The benefits and disbenefits of other materials, such as composts and anaerobic digestates derived from source segregated organic wastes, and untreated wastes spread under the current exemption regime, are less well understood, and the amount of characterisation data available is more limited.

As the UK endeavours to meet its obligations for renewable energy, reduction of greenhouse gases (GHG) and landfill diversion targets, the treatment of organic wastes and the types and characteristics of such treated wastes is likely to change. More organic wastes may be recycled to land that will have undergone some form of pre-treatment (principally either composting or anaerobic digestion). Where these new wastes have an agronomic benefit and are applied to agricultural land it will be vital that this has no adverse implications on the land bank available to accept these materials.

The impact of large scale changes in the quantities and diverse nature of organic wastes being recycled to soil needs to be carefully considered. Soil is a complex medium and its ability to maintain ecological processes, functions, biodiversity and productivity is vital to a sustainable future. A key rationale for recycling organic wastes to land is to supply nutrients for crop plant growth, reducing the need for inorganic fertilisers, thereby conserving raw materials and energy. The organic matter contained within the wastes may also improve soil properties such as soil structure and water holding capacity, reducing erosion and drought problems.

In this project data on the characteristics of a wide range of organic wastes have been collected and the agronomic and environmental impacts of applying these to agricultural land have been reviewed. This has been carried out through the development of a comparative methodology based on the known benefits of the organic materials.

A1.2 Objectives

Five specific objectives were identified to gain a better understanding of what is required to guide policy decisions in this area.

To provide an assessment of the pollution risk, greenhouse gas emissions and agronomic benefits from a range of organic materials recycled to land.

To recommend future research needed to develop guidelines and to assess trade-offs between nutrient recycling, losses to water and losses to air.

To recommend changes to guidelines for application to land of the above materials, if applicable.


4

To recommend suitable rates of application of anaerobic digestate to land, in order to optimise greenhouse gas (GHG) mitigation of the process while recovering nutrients.

To provide a short executive summary for policy makers.

A1.3 Background

A1.3.1 Principles of fertiliser addition to soils

Most UK agricultural soils require addition of the main plant nutrients N, P and K as fertilisers in order to optimize crops yields. Some soils also require additional nutrients such as S and Mg. The aim of fertiliser addition is to provide sufficient nutrients to match the crop requirements. Recycling organic materials to land may supply some of the fertiliser requirements cost-effectively and in a sustainable manner as it reduces the need for manufactured fertilisers. However, too much nutrient application may be a source of pollution, e.g. to ground and surface water from erosion and leaching of nutrients, and emissions to air from NH3 and N2O. Undesirable contaminants in organic wastes may also accumulate in the soil such as heavy metals.

Guidance on applying organic materials to soil as fertilisers at loadings that minimise environmental risks are described in the fertiliser recommendations for agricultural and horticultural crops RB209 (MAFF 2000 – henceforth referred to as RB209). Soil is a source of most plant nutrients and these may be present as soluble nutrients in the soil water, solid inorganic salts, adsorbed ions on soil particles and as part of the organic material of the soil. The amount available to the plant is usually considered as the soluble material plus nutrients that readily dissolve or mineralise from solid forms. However, these same soluble nutrient forms are also those most easily lost from the soil by leaching or metabolised by soil micro-organisms, leading to pollution of air and waters.

Standard methods (MAFF 1986) exist for assessing the amount of these readily-available nutrients and soils are categorised by indices for the various nutrients. RB209 provides guidance on the amount of additional fertiliser to add on top of the existing soil readily-available nutrient for different soil types and crops.

A1.3.2 General benefit of organic fertilisers

Organic fertilisers are usually considered to provide useful benefits for managing soil nutrients and improve soil structure in the agricultural environment. Many of the nutrients are present in the organic matter as minerals and are immediately available to the crop. Others are part of the organic matter and are mineralised and made available to the crop as the organic material decomposes. This may provide a slow release source of fertiliser throughout the growing season for the first crop after application when most decomposition is likely to occur. This is seen as advantageous as excessive nutrient concentrations are minimised and this limits potential adverse environmental impacts. Following the first year, the remaining organic matter is much more stable and effectively forms part of the total soil organic carbon (SOC). This organic matter is not totally recalcitrant but very slowly decomposes and acts as a reservoir of nutrients that contributes to the background basal soil nutrient content as measured by the soil indices. Therefore organic fertilisers affect soil nutrients in the


5

short, medium and long term whereas inorganic fertilisers may only impact in the short term.

Table 1.1 shows typical inorganic fertiliser application to soils with and without manure applications showing the reduction in inorganic fertilisers used for many crops. This table illustrates the important point that organic fertilisers rarely if ever totally replace inorganic fertilisers but are used in conjunction as a soil nutrient management package.

Table 1-1 Typical application rate of inorganic fertiliser to soils (kg/ha)

With manure Without manure

Crop N P2O5 K2O N P2O5 K2O

Winter wheat 168 21 36 194 33 40

Winter barley 108 34 51 141 36 49

Spring barley 94 41 46 103 36 62

Main crop potatoes 109 91 144 144 151 230

Winter oilseed rape 181 19 39 191 32 37

Sugarbeet 791 11 81 99 50 110

Mean overall application rates (kg/ha) of fertiliser nutrients according to manure usage in Great Britain 2007 (BSFP 2008).

A1.4 Approach

For this study a significant database of the characteristics of organic materials currently (or that may potentially be) applied to land for agronomic benefit has been collected. The data reflecting current waste characterisation requirements focuses on analysis of the main potential crop nutrients in the waste (N, P, K, Mg and S) and some potential contaminants, e.g. heavy metals. The data has been collected from field trials and published reports, as well as data submitted to fulfil regulatory requirements for spreading wastes to land. Organic materials differ considerably in their moisture and organic matter contents. In this study a methodology has been developed whereby normalised application rates of the organic matter to soils in terms of their N nutrient content and stable organic carbon content are used. Following normalisation of loads, the loadings of other parameters to soil were reviewed and compared to provide an assessment of the potential agronomic benefits and environmental risks posed by specific organic wastes.

Where the data was sparse, as was the case for a number of waste types and waste characteristics, the information was supplemented through review of the literature for qualitative information. In these cases the assessment provides an indicative view that may be refined as real data becomes available.

A1.5 Appendix Structure

This appendix contains the following sections:


6

Section A2 – Describes the waste types and waste treatments for which characterisation data has been collected. This section also discusses the agricultural benefit and potential environmental impacts of different waste characteristics and illustrates this with a description of the more common organic wastes typically applied to agricultural soil.

Section A3 – Describes the methodology developed for normalising waste application rates to soil and for assessing the loadings of other waste parameters.

Section A4 – Describes the results of the assessment when the wastes are applied as a N fertiliser.

Section A5 – Describes the results of the assessment when the wastes are applied as a soil conditioner.

Section A6 – Describes the results of the assessment of gaseous emissions from soil from the application of the organic wastes

Section A7 – Discusses the assessment of organic pollutants applied to soils from organic wastes.

Section A8 – Discusses the risk to ground and surface water posed by applying organic materials to land.

Section A8 – Discusses the risks from pathogens applied to soil from organic wastes.

Section A9 – Presents key findings and conclusions from this study.


7

A2. DATA COLLECTION AND RELEVANT WASTE CHARACTERISTICS

A2.1 Waste categories

Historically many wastes have been applied to agricultural land, many in an untreated form. The implementation of policy and legislation has significantly changed the way organic materials are managed. In future most organic wastes applied to land will have undergone some form of pre-treatment. Table 2.1 shows the usage of principal generic wastes with potential for future land application potential.

Table 2-1 Generic organic wastes types applied to soil

Waste type N C Use

Farmyard Manures

High readily available N

Organic C not fully stabilised

Fertiliser and soil conditioner

Slurry Manure High readily available N


Fertiliser

Sewage sludge High readily available N


Fertiliser

Digestates High readily available N


Fertiliser

Composts Low readily available N

Organic C fully stabilised

Soil conditioner

Sources of materials composted and anaerobically digested in the future may be quite diverse and therefore the characteristics of composts and digestates may vary considerably. Data was also collected for some waste types that have undergone different treatments. For example, sewage sludge may be treated in several ways prior to spreading to land. These include liming, anaerobic digestion under mesophilic or thermophilic conditions, and heat drying. Also sewage sludge may be treated by a combination of treatments such as anaerobic digestion followed by liming of the digestate.

Such diverse wastes and treatments do not always readily allow division into existing waste classification systems. Therefore for this study a bespoke classification system has been used which provides logical grouping of collected waste type data (Table 2.2) and waste treatment data (Table 2.3). The groupings are then colour coded for easy recognition within the graphical representations of the data. These groupings aid and illustrate the evaluation methodology. The methodology could be applied to assess any waste type on a case-by-case scenario as it focuses on the waste composition rather than any particular classification of the waste.


8

Table 2-2 Waste types and colours used in this study

Waste class Description

10 Manures

11 Cattle slurry

12 Pig slurry

13 Cattle manure

14 Pig manure

15 Sheep manure

16 Poultry manure

17 Horse Manure

18 Livestock slurry

19 Manures (general)

20 Waste from waste water treatment

21 Screenings

23 Sludges from biological treatment of industrial waste

24 Activated sludge

25 Biological aerated filters

26 Sewage sludge (general)

30 Greenwaste

31 Plant tissue waste

32 Garden and park waste

33 Seperately collected fraction MSW (curb side collections)

34 CA site greenwaste

35 Grass cuttings

36 Source segregated green waste

37 Wood

38 Greenwaste (general)

40 Foodwaste

40.1 Dairy production waste

40.2 Catering waste

40.3 Kitchen waste

40.4 ABPR waste

40.5 Foodwaste (general)

41 Sludges from washing and cleaning for food preparation

42 Gut content


9

Waste class Description

43 Blood

44 Vegetable washings

45 Vegetable production waste (peelings, choppings)

46 Sugar processing

47 Wastes from baking and confectionairy

48 Beverage production

49 Fish farm waste (faeces and uneaten food)

50 Biowaste (mixed organic wastes)

50.6 Green waste and COM

50.7 MSW and manure

51 Bio waste municipal

52 Biodegradable kitchen and canteen wastes

53 BMW

54 MBT residues

55 Manure and biobin material

70 Others

71 De-inking sludges from paper recycling

72 Wastes from processing sheep fleeces

73 Dredgings

74 Construction and demolition wastes (soil)

75 Chipboard

76 MDF

77 Cardboard

78 Sludges from treatment of drinking water

N Not specified


10

Taebl 2.3 gives the treatment type codes used in this study, together with the predicted biodegradation rate (see Section A3), and the percentage ammonia lost to atmosphere (see Section A6).

Table 2-3 Waste treatment codes used in this study

Code Waste Type

Biodeg. NH3 emission

% C min % NH3

CL Composted and lime 10 50.0

CM Composted mechanical - mechanical dewatering 10 50.0

C Compost 10 50.0

DM Meso AD and mechanical dewatered 60 50.0

D Mesophilic AD 60 36.5

H Heat dried 60 50.0

LS Liquid storage 60 50.0

L Lime stabilised 60 43.3

M Mechanical dewatered 60 50.0

N Not specified 60 50.0

Q Liquid fraction 60 24.9

RL AD + Lime stabilisation 60 48.4

R AD 60 50.0

TM Thermo AD and mechanical dewat 60 50.0

T Thermophilic AD 60 50.0

U Untreated 50 38.9

A2.2 Waste characteristics considered

A2.2.1 Waste parameters for normalisation of application rates

In the agronomic assessment of wastes, application rates to soil have been assessed in terms of using the organic waste as either a fertiliser or soil conditioner. Application rates for these two scenarios have been normalised based on the following to allow a comparison of the benefits and disbenefits of the wastes.

Potential available nitrogen (PAN) – the waste application required to give the same loading of PAN at 100 kg N/ha.

Soil organic carbon increase – the waste application required to increase the SOC by 10,000 mg C/kg in the top 25 cm of soil.


11

A2.2.2 Waste parameters collated from available data

The waste characterisation data sought and considered relevant to this study were:

Dry matter content;

pH;

Organic matter- loss on ignition (LOI), total organic carbon (TOC) and dissolved organic carbon (DOC) contents;

Nitrogen content – total (TN) and mineral (RAN);

P, K, Mg, S content – total and available;

Metals – total Cd, Cr, Cu, Ni, Hg, Pb and Zn;

Pathogens – plant, animal and human pathogens;

Organic pollutants –persistent toxic organic pollutants;

Biodegradability;

Leaching characteristics. For many wastes it was not possible to collect data for all these parameters and where this was important for the assessment assumed values were used which can be updated when more reliable data becomes available.

A2.3 Waste characteristics background

The parameters are discussed in the following section as introduction to their impact as either nutrients and/or environmental risk.

A2.3.1 Dry matter content

Wastes applied to land may vary in moisture content, from being liquids to virtually completely dried solids. Within any waste group there may also be a significant variation in moisture content. Application rates of wastes to land are often reported as wet weights and this therefore includes the moisture content. Normalisation of the application rates for this study requires that the loadings are expressed on a dry matter basis. Fortunately most waste characterisation includes determination of the moisture content.

A2.3.2 Organic matter content

The organic matter content of wastes is often measured as either the loss on ignition (LOI) or as the total organic carbon (TOC) content. This is a key parameter for assessing the impact on soil organic carbon (SOC) content and potential release of nutrients from the organic matter, when used in combination with the biodegradability characteristics of the organic matter, as it decomposes in soil. Carbon data has been used for the assessments and therefore used the relationship of TOC = LOI/1.83 to convert LOI data to TOC data, i.e. this assumes the LOI is 54.6% organic carbon.

The dissolved organic carbon (DOC) in the organic wastes is a sub-fraction of the TOC and may be readily leached and contaminate groundwater and surface water. Indicative calculations show that this may be substantial for some wastes and that the material may be either recalcitrant in stabilised wastes such as composts and therefore persist in the environment, or readily biodegradable in untreated wastes and inhibitory to plant growth. The DOC may also complex with metals and increase


12

leaching of these materials (Comans et al. 1993), and may also comprise specific toxic organic pollutants themselves.

A2.3.3 pH

The pH of the wastes may be important in controlling the availability of some nutrients by affecting their solubility. This is particularly relevant for wastes with high mineral NH3 contents where volatilisation of NH3 from the waste during application may be significant. These issues are discussed further in Section A3.4.

A2.3.4 Total nitrogen, organic nitrogen, nitrate and ammonium

The waste N content may be present in several forms, and the relative level influences the value of the waste as a fertiliser and its environmental risks.

Total nitrogen (TN) – is a measure of the total N content of the waste and includes all the mineral and organic nitrogen species.

Mineral nitrogen – this is the N in the waste present in mineral form as NH3, urea, uric acid and nitrate. Plants only use nitrogen as fertiliser when it is in the form of NH3 and nitrate. Urea and uric acid however, readily decompose to NH3. The mineral nitrogen is often described as the readily available nitrogen (RAN). Note that in this study the term ammonia (NH3) has been used for the total ammoniacal nitrogen (NH3 + NH4

+) content of the waste.

Organic nitrogen (ON) – this is the nitrogen found in organic nitrogenous compounds such as proteins, amino acids and nucleic acids. During decomposition of these materials this N is mineralised to NH3. It is important to note that some organic nitrogen may be locked up in stable organic nitrogenous compounds that do not decompose. This includes being locked up in new microbial biomass following growth of the micro-organisms on the waste. Not all the organic N is therefore considered available as plant nutrient and the amount will vary from waste to waste. In general the organic N in fully stabilised composted wastes is not a good source of fertiliser N.

The fertiliser recommendations handbook: RB209 provide guidance on the amount of fertiliser N required for specific crops. For organic wastes this is defined by the potentially available nitrogen (PAN), although fertiliser recommendations may also refer to crop available nitrogen (CAN) which is the PAN less any losses of PAN from leaching, microbial nitrification and denitrification, and volatilisation. The PAN represents the sum of the mineral nitrogen (RAN) and the amount of ON that may be released as mineral N during the initial waste decomposition in the soil. The PAN therefore represents the potential nitrogen fertiliser value of the waste to the first crop.

However as some wastes may contain ON that is relatively stable and not made available to the first crop as PAN. This ON is added to the bulk soil N content and may over many years very slowly be mineralised contributing to the basal soil nitrogen supply (SNS). Ammonia derived from the decomposition of the soil organic N is not very mobile in soils as it binds to cation exchange sites and is not considered a leaching problem. However, most NH3 in soils is rapidly converted to NO3

- following microbial nitrification. Nitrate is very mobile and if the SNS is high and crop growth is not occurring the NO3

- has the potential to leach to groundwater.


13

In topsoils typically more than 90 – 95% of the soil N is present as organic N, most of which is present as amino acids, amino sugars and heterocyclic ring structures all in high molecular weight polymers (Russell 1988). Soils also have C/N ratios of between 10 -14 which for a typical sandy soil with an SOC of 1% gives an N content is about 2.5 t N/ha (833 mg N/kg soil DM). Higher N concentrations would be expected in soils with greater SOC contents.

The mineralisable fraction of the ON is usually considered beneficial as it is released slowly as the organic waste decomposes. This provides a constant supply of fertiliser but at concentrations that do not pose an environmental risk from being present in excess. However, crops do not take up N in a linear manner throughout the growing season and maximum N uptake occurs during the period of leaf expansion. For cereals this will be during April and May, after which, N uptake is greatly reduced and growth maintained by re-mobilization of N already in the plant. Hence, the continued release of RAN from organic manures may have two impacts:

either the crop will take up more N than needed, leading to increased N in crop residues. This in turn may lead to greater losses of N2O and NO3

- when those residues break down in autumn, or

the RAN may accumulate in soil to be lost as NO3- by leaching over winter.

Hence pre-treatments that decompose organic matter and retain the mineralised N as RAN, e.g. anaerobic digestion, may improve crop uptake of N and reduce subsequent losses if applied at appropriate times.

Potentially available nitrogen (PAN) - Describes the sum of the RAN and the amount of organic N that may be mineralised in a relatively short time period (first year following application).

A2.3.5 Phosphorus (P)

General principles

Phosphorus is an essential plant nutrient and can be supplied as fertiliser in organic materials. In soils, P is found mainly as phosphates, which may be in either an inorganic or organic form. The majority of the P is present in solid form and as some of this may be relatively stable only a fraction of the total soil P may be available as a plant nutrient. The readily available plant nutrient P content of soils comprises soluble forms and the types of solid forms that most readily dissolve. This fraction is often measured as a soil P index (RB209) where the available P is determined as either the Olsen-P (P extracted in sodium bicarbonate solution, MAFF 1986) or resin P. For most crops a soil P index of 2-3 is suitable. At higher P indexes there is an increased risk of adverse environmental impacts from the excess available P (Withers et al. 2001). The available P in the soil is replenished by the dissolution of the less stable solid inorganic phosphate forms and the decomposition of organic phosphate.

Phosphate leaching to surface waters is a prime cause of pollution and is of high concern. Such pollution of surface waters may also occur from erosion of soil where much more of the P content of the soil entering the water course may dissolve in the much larger volume of water. As surface waters are generally limited by P, small concentration increases may have significant adverse effects from eutrophication.


14

Organic fertilisers will contain a similar mix of organic and inorganic P forms with differing degrees of stability. The P from organic waste applications may then contribute to the available P in the soil as a fertiliser although it may do this through a slow release process that in part depends on the decomposition of organic P sources. In RB209 action of the total P in organic fertilisers is assumed to be available (60% for manures, and 50% for slurries and sewage sludge). There is however some debate on the actual availability and therefore fertiliser value and environmental risks posed by the P added to soil in organic materials. The P content of organic wastes may be measured as total P (BS EN 13650), water soluble (BS EN 13652) and DTPA soluble (BS EN 13651) as described in PAS100 composts as well as by the Olsen-P method. Typically a water extraction gives a lower extractable P than the DTPA method. The Olsen-P method is similar to the DTPA method.

Generally sewage sludge is rich in P. Therefore when applied as N fertilisers significant loads of P are applied, which may exceed crop requirements. The actual availability of the P applied in sewage has therefore come under scrutiny and a recent study UKWIR (2006) suggested that the available P in sewage sludge may be lower than the values assumed in RB209 and consequently of less environmental risk. The data indicated that the mean Olsen-P in 7 sludges was about 9.8% of the total P. Incubation of several soils for 90 days with an equivalent P loading of 200 kg/ha, the soil increase in Olsen-P was equivalent to a mean of 8.3% of the total P added, i.e. the same as the Olsen-P in the sewage sludge. This implies that the remaining P in the sludges was completely stable.

In natural soils there is continuous removal of available P by crops and by leaching and therefore an equilibrium is not established between P in solution and solid P forms. Extraction methods such as Olen-P may be equilibrium based and underestimate the real potential available P. Also the soils were air dried after incubation which may also have resulted in some available P being rendered unavailable for extraction.

The acid ammonium oxalate extraction method is often applied to inorganic fertilisers and soils. It will extract more of the insoluble phosphate than Olsen method. As much of the P in soils and organic wastes may be in these solid inorganic forms it may be a better measure of P availability than Olen-P method.

Elliott & O‟Conner (2007) recognized that the plant P availability of sewage can vary but that most are in the range of about 25 – 75% of the availability of an equivalent amount of inorganic fertiliser (triple super phosphate TSP). Higher P availability (>75%) was indicated for biological P removal process derived sludge (BPR) and low availability (<10%) for high Fe, heat dried and lime stabilised sludge. The P availability in a generic waste type may therefore be quite variable and depend on the exact waste source and its pre-treatment. Eghball et al. (2003) showed organic P was 16% of total P in composted manures and 25% in non-composted manures, and up to 75% in both was inorganic phosphate. Most of the inorganic phosphate was assumed to be present as Ca-phosphates which are not very soluble in water or bicarbonate, i.e. not readily extracted the Olsen-P method.

Organic P content of soil may vary widely and constitutes between 20 and 80% of total soil P and must be mineralised to become plant available. Oehl et al. (2004) showed that organic P mineralization constituted only about 10% of the rate of P exchange between solid and soluble forms indicating that physico-chemical processes are more important than organic P mineralization for releasing available phosphate. The impact and risk from these P species in soil if soil is eroded into


15

surface waters needs some consideration, as solubilisation in a virtual unlimited water supply may mean a significant dissolution of the soil inorganic P content.

Repeated N-based applications of manure leads to an accumulation of soil P in excess of crop needs which can then increase P enrichment of agricultural runoff (Sharpley and Moyer et al. 2000). The ratio of PAN:P in manure derived compost is significantly less than the ratio in the un-composted manure (Preusch et al. 2002, Sikora and Enkiri 2000) and less than the ratio of N:P in crops (Mullins and Hansen 2006). Therefore the accumulation of P in soil is likely to occur when some composts are repeatedly applied to supply crop N needs.

The application of biosolids at rates based on crop N requirements usually has P in excess of crop needs. Elliot et al. (2002) conducted laboratory and greenhouse column studies to characterize the phosphorus leaching and the forms of leached phosphorus, following the application of biosolids, compost and inorganic fertilizers on sandy soil. Bahiagrass was cultivated in soils with low phosphorus content and low phosphorus sorption capacity in the columns. The leached phosphorus was predominantly inorganic, and was in lesser concentrations for biosolids than for inorganic fertilizers. The percentage of applied P leached varied from 1.7% to 21.7% from the inorganic fertilizer treatment. The presence of phosphorus in the leachate from biosolids and compost amended soils was < 1% of the applied phosphorus. Esteller et al. (2009) found that % P leached was less than 1% for all the treatments, values similar to those obtained by Elliot et al. (2002). These results indicate that P availability in organic wastes may be less than inorganic P fertilisers.

Iglesias-Jimenez and Alvarez (1993) studied the effect of city refuse compost as a P source to overcome the P-fixation capacity of sesquioxides-rich soils. They showed it effectively diminished the fixation process by providing equivalent amounts of soil labile-P as di-potassium hydrogen orthophosphate, which significantly increased P concentration of plant tissue. Hue et al. (1994) also reported similar findings using yard-waste compost and attributed this to the release of P during the decay process and the competition between organic anions (released by compost) and P for adsorption sites in the soil complex.

The view is taken that it is not clear how much P is available in different organic wastes and that it is possible that individual wastes will vary significantly depending on the exact source and treatment of the waste. An appropriate waste characterisation method is not available. It is premature to provide assurance that waste characterisation based on Olsen-P or similar provides a reliable prediction of the P risks. A conservative view is taken and it is assumed that a significant fraction of the total P is available.

A2.3.6 Potassium (K)

Potassium is an essential plant nutrient and is often provided in inorganic fertilisers. It is ubiquitous in the natural environment and unlikely to present a significant environmental risk of causing harm if excess K is added in organic wastes. Therefore only beneficial aspects as a fertiliser have been considered.

Potassium is very soluble in water and therefore virtually all the K in any fertiliser is usually considered as bioavailable. In RB209 the available K is taken as 90% of the total K content for manures and sewage sludge. Applications of K fertiliser are guided in RB209 by assessing the soil K content as an index in an analogous methodology to that used for P and then applying sufficient additional K for crop requirements.


16

A2.3.7 Sulphur (S)

Sulphur is another key major plant nutrient and in RB209 the S nutrient quality of organic fertilisers is based on the total S content. Whilst there are some soils where sulphur deficiency occur this is easily remedied by the application of inorganic fertiliser. Sulphur as sulphate is ubiquitous in the environment and unlikely to cause significant environmental harm if supplied in excess as inorganic salts under normal aerobic soil conditions. However S may exist in several forms which may be metabolised by microbes in soil.

In organic wastes S may be present as sulphates and as organic S in proteins. It may also be present as inorganic sulphides in anaerobic digestates produced by the action of sulphate reducing bacteria, mostly lost as hydrogen sulphide in the digestion process. Reduced sulphur compounds can be very odorous and soluble sulphide is also toxic and may be released as the very toxic and odorous hydrogen sulphide gas under acidic conditions. Sulphides may also form insoluble metal sulphides reducing the bioavailability of trace nutrients.

Sulphide production may occur in soils through the action of sulphate-reducing bacteria if the soils are anoxic (water-logged) for any length of time and contain sulphate combined with a high level of biodegradable organic matter (high biological oxygen demand, BOD). Such conditions may occur following addition of unstabilised organic wastes to soils where sulphate is already present in the soil or is added within the organic waste.

Because of the complexity of the biology of S it may present some risk of environmental harm under some circumstances and therefore require greater control on applications compared with K.

A2.3.8 Magnesium (Mg)

Magnesium is also an important plant nutrient and RB209 provides guidance on application of Mg fertiliser. As with P and K this is based on determining the soil Mg content and applying additional Mg to meet crop requirements. For manures and sewage RB209 assumes all the Mg in the waste is available as plant nutrient. As there is little environmental risk from Mg or its overdosing, then, as in the case of K, only the benefits need to be considered.

A2.3.9 Metals

Several heavy metals are present in sewage sludge and have been the source of much research on the environmental impacts of the build up of heavy metals in soils. Sewage sludge applications to agricultural land are regulated in England through the Sewage Sludge Regulations (SI 1989) with additional guidance in the Code of Practice (Defra 2003). These regulations and guidance contain limits on the concentration of particular metals (Cd, Cr, Cu, Hg, Ni, Pb, Zn and Mo) in the soil and the rate of metal application as a 10 year average. Therefore regulation is based more on the soil metal content and soil protection than on the actual metal content of the sewage sludge

Although several of these metals are also essential plant trace nutrients (e.g. Zn, Cu, Ni, Mo) it is rare that soils are deficient in these trace nutrients. Where deficiencies occur it is usually a result of the availability of the metals for plant uptake being


17

restricted either by pH or soil organic matter. Copper deficiency does occur occasionally in some soils and Cu is a key metal in the enzyme nitrous oxide reductase responsible for N2O reduction to nitrogen gas during denitrification and the enzyme methane monooxygenase responsible for CH4 oxidation. Hence most organic additions to soil would be monitored in terms of limiting the build up of heavy metals in the soils. Metals such as Cd, Pb and Hg have no known biological function and pose a risk of causing environmental harm to sensitive organisms in the soil environment and of entering the human food chain through uptake into crops.

Other important factors relevant to soil metals are the bioavailability e.g. for uptake into food crops (Peralta-Videa et al. 2009) and the leachability of the metals in the soil. The bioavailability and mobility of metals in soils is complex and dependent on several factors including the soil physico-chemical conditions and the form of the metals in question. The impact of high soil metals on soil function especially microbial activity has been the subject of much research. In Defra/UKWIR sponsored long term trials on the effect of metals from sewage sludge applications on soil quality (Chambers et al. 2003) there was a reduction in soil Rhizobia numbers in the high Zn amended sewage sludge plots. The Rhizobia numbers were stable in plots amended with only Zn salts indicating that the response was not due to Zn alone. In these trials high loadings of organic matter were applied and it is conceivable that the response was due to the combination of the organic matter decomposition enhancing the impact of Zn on the Rhizobia. This hints at potential interactions between different waste characteristics on soil responses.

Another key factor in determining the form and concentration of metals is the pre-treatment received by the organic waste prior to land-spreading. Generally composting processes result in a decrease in dry matter content and consequent enrichment in the conserved metals. However, the leachability, and by inference the bioavailability of the metals may have been reduced by the pre-treatment. Therefore the composted material would generally be considered to pose a lower risk, although whether this is a permanent response is a matter of debate (Godley et al. 2008). However there may be exceptions as Miaomioa et al.. (2009) showed that the availability of Cu and Zn increased when composting sewage sludge.

The addition of organic wastes may increase the amount of leachable metals or enhance the leachability of metals already in the soil, e.g. by adsorption to DOC (Comans et al. 1993).

Weber et al. (2007) monitored a sandy soil for three years following addition of two MSW derived composts at different application rates of 18, 36 and 72 t DM/ha. One of these composts was from an industrial centre and had high metal contamination. The highest application rates of this compost produced significant increases in soil metal concentrations for Zn, Pb, and Cu. Whilst the soil Pb content remained stable, there was a decline in the Zn and Cu contents. This is indicative of significant leaching of these metals from the soil despite the water soluble metal content being only a small fraction of the total metal content in the waste.

A2.3.10 Organic waste biodegradability and stability

Organic wastes applied to agricultural soils will have been pre-treated by a variety of methods which may include heat drying, liming, anaerobic digestion and composting, and consequently will vary in their biodegradable organic matter content. This will affect its behaviour in soil especially with respect to the release of plant nutrients and potential environmental impact.


18

The biodegradation of organic matter in soils provides the main energy input into the soil to sustain microbial life. Therefore it may be considered a beneficial practice for the soil ecology to apply organic matter that is not fully stabilised and that decomposes to some extent. However, the application of large amounts of organic matter into soils in a single application is not a natural occurrence as most organic carbon inputs are derived from plant root exudates and turnover of roots and foliage material over a period of time. Agricultural practices such spreading livestock manures, ploughing in of grass and crop residues such as sugar beet tops represent the unnatural application of significant amounts of organic matter to soil in a single event. Moreover the amounts of stabilised organic wastes added to land to significantly increase SOM may add greater amounts of dry matter than currently accepted practices.

Two biodegradation properties of the organic waste may be of importance when assigning risks and benefits. The first is the extent of decomposition achieved in the soil. For most organic wastes applied to soils the majority of the organic matter decomposition occurs in the first year of application. This corresponds to the decomposition of the readily and slowly biodegradable materials in the organic matter. After the first year the remaining waste derived organic matter would be expected to consist mainly of very poorly biodegradable organic matter and decompose very slowly. For fully stabilised composts however, the majority of the organic matter would persist and be slowly degraded over many years with little decomposition expected in the first year.

Many studies have monitored the decomposition of organic wastes in soils which confirm this principle. For example, Bernal et al. (1998) composted several mixtures of organic wastes to full maturity then monitored CO2-C evolution from the untreated, intermediate (at the peak and after the initial active composting phases), and fully composted samples in soil over a 70 day period. The percentage of the waste organic carbon released as CO2-C decreased with increasing composting time (Table 2.4).

Table 2-4 Percentage carbon mineralised in soils after 70 days incubation for wastes composted for different times

Compost feedstock

Untreated At peak of active

composting

At end of active

composting

After full maturation

Sewage sludge and cotton waste

62.3 37.7 22.4 19.8

Poultry manure and cotton waste

68.2 43.8 28.6 24.3

Sewage sludge and maize straw

93.2 77.4 40.1 23.7

Pig slurry and Poultry manure and sorghum

bagasse

38 14.7 13.1

(Bernal et al. 1998)


19

The second important biodegradability factor is the initial rate of decomposition. The decomposition of most organic wastes is most rapid at the start when the decomposition of the most readily biodegradable fraction of the waste occurs. During this period the waste will exhibit its highest biological oxygen demand (BOD). If a waste has a high initial BOD and is applied at a significant loading, the BOD in the soil may be too high and the soil become anoxic. Such conditions may increase the rate of N2O emission through denitrification as well as cause stress to aerobic soil organisms affecting soil function.

Organic matter (OM) degradability is a key concept in soil OM studies and can be defined as the capacity of OM to be utilized by soil microbes as a source of energy (Rovira and Vallejo 2002). Several aerobic waste biodegradation tests are available which monitor the initial biodegradation rate (Wagland et al. 2009). Alternatively soil respiration testing may be carried out immediately after the application of the waste to monitor the initial rapid degradation period (Mondini et al. 2008). Development of appropriate correlations between these approaches may allow prediction the impact of the waste on the soil BOD from laboratory waste biodegradation tests. Many other parameters and indexes have been proposed to measure OM degradability, e.g., C/N, lignin/N, and cellulose/lignin/N (Rovira and Vallejo 2002). Organic matter composition with respect to macromolecular components appear to be more important than the other parameters in determining degradability. For example, lignin, phenol, and tannin contents in the OM were more important in determining soil OM decomposition than C/N ratio (Paustian et al. 1997). The partition of compost OM chemically into recalcitrant and labile pools has been shown to correspond to carbon retained and degraded after soil incubation (Rovira and Vallejo 2002, Adani and Ricca 2004).

The response of the soil to different degrees of stabilization of biosolids was studied by Sánchez-Monedero et al. (2004) who found an increase in the activity of the microbial biomass in the soil which was related to the level of stabilization of the composted biosolids. Thus, application to the land of stabilized products reduces the extent of disruption to the microbial biomass in the soil. The initial increase in microbial biomass in the soil after the application of organic matter was related both to the readily biodegradable C input for the original soil microbial biomass and to the microbial biomass already growing in the material added to the soil (Perucci 1992).

Waste stabilization by means of anaerobic digestion reduced the amount of readily-biodegradable materials (Sánchez-Monedero et al. 2004). Only a short period, between 1.2 and 1.4 d, was needed to break down half of the labile-C fraction of the wastes. It was estimated that 25% of the C in cattle manure digestate that was capable of mineralization was calculated to be readily mineralized, while the rest (the recalcitrant fraction) would require 94 d to be reduced by half. In the case of the poultry manure digestate, which was less mineralized, 47% of the C was mineralized slowly, needing 74 d to reduce its content to a half. This result was in agreement with the greater amount of material remaining after thermal analysis (Sánchez-Monedero et al. 2004). The C fraction readily broken down by microorganisms was represented by mass losses at low temperatures (300°C) in differential thermogravimetry (DTG) profiles. Agreement was reported between the results from maximum velocity of oxygen up-take (VOU) (from biodegradability test) and the intensities of such peaks. In this way, the cattle manure digestate, with its greater VOU value, also had a higher intensity peak in the DTG profile than did the poultry digestate (Sánchez et al. 2008).

Marcato et al. (2008) studied changes in pig slurry organic matter (OM) during anaerobic digestion (AD) in a reactor to characterize organic matter (OM) evolution through AD. Organic matter maturity and stability were evaluated using different


20

biological and physico-chemical methods. Seed germination and growth chamber experiments revealed a greater maturity of digested slurry (DS) than raw slurry (RS). Soil incubations showed that DS was more stable than RS with a C-mineralization of 12.0 g CO2-C 100 g-1 Corg after 49 days, as compared with 17.6 g CO2-C 100 g-1 Corg for RS. Biochemical fractionation showed a relative increase in stable compounds such as hemicellulose-like and lignin-like molecules. Fourier-transform infrared spectroscopy showed some changes in the chemical structures of OM with a reduction in the aliphatic chain, lipid and polysaccharide concentrations. A comparison between the evolution of OM during AD and the first weeks of a composting process showed almost identical changes. Finally a theoretical method called fictitious stomic-group separation (FAS) was applied to the elemental compositions of RS and DS. DS was less humified than RS and presented the properties of a fulvic acid, indicating that the observed stability in DS was mainly due to the biodegradation of the most labile compounds.

These workers concluded that spreading stabilised anaerobically digested organic matter should be less disturbing for the soil microflora than spreading raw organic matter. Consequently, there will be less risk of competition for nutrient (N in particular) and oxygen between the crops and the soil bacteria.

Soils are an important store of fixed carbon. Smith et al. (2000), discussing the potential of C sequestration in Europe, pointed out the lack of data on compost management as a means of C sequestration. Barral et al. (2009) calculated that, between May and October while maize was growing, following compost applications of 50 and 85 Mg ha-1, 369 g C kg-1 (40%), and 520 g C kg-1 (53%) of the C added was retained. Relative C degradations (g C kg-1 compost-TOC) for both treatments were not much different from the relative content of labile C in compost. This was 414 g C kg-1 TOC, calculated by using data on relative C contents of the labile and recalcitrant fractions (g kg-1 OMlab or OMrec) and the relative fraction contents of compost (g kg-1 OMcomp; Table 2). Therefore, these results suggest that, after 150 d of compost application, the labile fraction was completely degraded and the recalcitrant fraction had not been degraded. The non-hydrolyzable OM (biochemically protected OM), i.e., the recalcitrant fraction was previously shown to be the fraction that contributes to the stable OM of soil (Rovira and Vallejo 2002; Mikutta et al. 2006). These data point to a slight increase in the fraction of C retained with increasing compost application. This is in contrast with literature data that showed constant or decreasing percentage of C accumulation with increasing amount of OM application (Körschens and Müller 1996; Smith et al. 2000). One possible explanation is that the compost used in this experiment was highly stable with a recalcitrant fraction over 50%, and other experiments may have used more degradable compost.

The authors concluded that compost use represents an interesting opportunity to sequester C in the soil, especially in the light of increasing interest on agricultural compost use. As compost represents 'predigested' OM, the large content of recalcitrant-OM allows for substantial soil-C accumulation levels after a short time. In this study, both the amount of carbon retained and degraded increased with the amount of compost applied. The fraction of the C added that was retained appeared to increase slightly with the increased compost dose. Further investigations are needed to better elucidate on how compost characterization in labile and recalcitrant C fractions can predict the C sequestered in the soil in short- and long-term experiments.

Barral et al. (2009) calculated the amounts of compost to be added to three soils every year to meet the needs of SOC under two scenarios. The first was the conservation of the current SOC concentrations, and the second one was to reach a


21

SOC concentration of 3.5%, that had been considered as a threshold value for the productive and structural functions of soils of temperate regions. The amounts of compost to be added were estimated to range between 4 and 10 t ha−1 (dry weight), depending on the targeted SOC values and the temperature. It was noted that an increment of 2°C in the mean temperature would increase by 10% the requirements for composts to maintain adequate SOC. Mineralization rates were obtained after incubation of the samples at 25°C for 10 weeks. Prior to incubation, soil moisture was adjusted to c. 70% of field capacity. 25 g samples of each soil, compost, or soil/compost mixture were placed in 100 ml incubation vessels, three replicates per soil/compost, and introduced in 1 l jars, each jar containing 25 ml of water, and 20 ml of 1N NaOH in a 50 ml vial. The jars were hermetically sealed and incubated at 25°C for 10 weeks. The CO2 evolved by microbial respiration was measured twice a week by titration of the residual NaOH with 0.1N HCl for the soils and 0.5N HCl for the composts. These data were transformed to percentage of total C evolved as CO2-C, and the cumulative mineralization data were fitted to first-order kinetic models. The authors concluded that the mineralization rate of SOC in field conditions can be adequately predicted from laboratory incubation experiments; the doses of organic amendments needed to counteract the losses due to mineralization or to reach fixed SOC objectives in agricultural soils can therefore be approximately estimated.

Organic matter applications to soil and its associated impact on soil microbiology may also significantly affect soil properties. The production of mucilages by bacteria and fungi has been found to enhance the formation of soil microaggregates (Oades 1993, Six et al. 2004). Tejada et al. (2009) reported the addition of composted plant residues also increased soil microbial activity and considered this could be responsible for the increase in the soil structural stability. However, the mucilages produced by bacteria and fungi represent a source of labile soil organic carbon (Oades 1993). Martens (2000) indicated that the aggregate binding effect of labile soil organic carbon is rapid but transient, while slower decomposing soil organic carbon has subtler effects on aggregation, but the effects may be longer lived. Aggregate stability increased more in the soils amended with composted plant residues of greater humic acid concentration. Soil bulk density was found to decrease during the experimental period as a result of dilution of the denser soil mineral fraction and soil aeration increased because of the increase in soil porosity accompanying structural stability. They considered soil microbial biomass and soil respiration were greater in the soils amended with composted plant residues with a greater fulvic acid concentration which may have been due to a greater labile fraction of organic matter in these residues.

Evanlyo et al. (2008) applied a compost derived from poultry litter and beef manure from a feedlot. The high compost rate treatments reduced soil bulk density compared with other treatments by the end of the second growing season. After 3 years of compost applications, even the smaller compost rates decreased bulk density. This demonstrated that continuous application of less than agronomic compost rates could improve soil physical properties of the Fauquier silty clay loam. The authors cite Khaleel et al. (1981) who, in a review of 14 research articles, found a direct correlation between decreasing bulk density and organic C additions. Soil porosity exhibited trends similar to bulk density. Differences in soil water-holding capacity were not distinguishable among treatments until the final year. The low compost treatments did not add enough organic material to increase the soil water-holding capacity.

Courtney and Mullen (2008) reported increases in soil organic matter (SOM) four months after application of MSW at 50 and 100 t ha-1. Significant reductions in bulk density were also measured following application of 100 t ha-1 – although this was


22

thought to be a dilution effect. Bhogal et al. (2006), in their study of the effects of additions of OC on soil physical properties, reported that large (65 t C ha-1) inputs of organic manures were needed to produce measurable changes in soil physical properties. Nevertheless, significant increases in total SOM and active SOM were measured following additions of manure, total SOM and active SOM increasing by c. 3% and 14% for each 10 t ha-1 addition of manure-C.

The manure inputs were associated with changes in topsoil porosity (+0.6%), plant-available water capacity (+2.5%), topsoil shear strength (-3.7%) and bulk density (-0.5%) for every 10 t ha-1 manure-C applied. The size and activity (as measured by respiration rate) of the microbial biomass were increased by 11 and 16% respectively for every 10 t ha-1 manure-C applied. No impacts were measured on aggregate stability as measured by the dispersion ratio.

These measured changes were consistent with the reported finding of Jenkinson (1988) that the proportion of total C remaining in the soil following organic manure applications is similar, once the initial rapid phase of decomposition has finished, from a range of organic inputs, with c. 30% of the C remaining in soil within one year of the application.

A2.4 Emissions to air

A2.4.1 Ammonia

Ammonia (NH3) emissions to air are considered to occur mainly during the application of the waste and are dependent on the waste characteristics, especially the NH3, pH and moisture content.

Estimates of ammonia emissions available from the literature

Moller and Stinner (2009) compared emissions from digested and undigested cattle slurry. They reported that cumulative NH3-N emissions after 12 h were significantly (P < 0.05) greater from digested than from undigested slurry, albeit the differences were not large at 10.2 and 8.9% of N applied respectively. For the remainder of the measurement period emission rates were equal. These results did not suggest that digestion significantly enhanced NH3 emissions of materials already rich in NH3. Hjorth et al. (2009) also reported no significant difference in NH3 emissions following land application of digested and undigested slurry. Rubaek et al. 1996 has reported fewer emissions from digested manures, whereas others found similar (Pain et al. 1990; Wulf et al. 2002a; Chantigny et al. 2004) or greater emissions (Amon et al. 2006))

Chantigny et al. (2007) attributed these reported inconsistencies with respect to the impacts of digestion on emissions of NH3 by the increase in pH of liquid manure during AD, which increases the potential for NH3 loss from digested products, being off set from reduction in dry matter content of digestates improving infiltration into the soil (Pain et al. 1990; Wulf et al. 2002a). Chantigny et al. (2007) reported AD to increase pH by almost one unit. This effect, noted in previous studies (Pain et al. 1990; Kirchmann and Lundvall 1993; Chantigny et al. 2004), was mainly attributed to the decomposition of volatile fatty acids into CH4 during the digestion process (Sommer and Husted 1995).


23

A general methodology has been developed for the estimation of NH3 emissions from manures and slurries (see Section A3.4). The anaerobic digestion of other wastes such as food wastes may be from a variety of AD process designs and post digestion treatment resulting in digestates with different pH, moisture and NH3 contents. This same approach has been used in this study for assessing potential NH3 emissions for such wastes (section A3.4).

A2.4.2 Nitrous oxide

Estimates of nitrous oxide emissions available from the literature

Some authors reported similar N2O emissions between treated and raw liquid manure for anaerobic digestion (Wulf et al. 2002b), whereas others reported decreased emissions (Rubaek et al. 1996, Petersen 1999, Vallejo et al. 2006) for treated manure. Wulf et al. (2002b) reported similar N2O emissions between anaerobically digested and raw manure when applied to arable soils but greater emissions with the digested manure when applied to grassland soils.

Chantigny et al. (2007) found digested slurry was generally associated with significantly less N2O emission than the other slurries. Losses were 54 - 69% less with the digested than with raw liquid swine manure (LSM) in the loam soil and 17 to 71% less in the sandy loam. This apparent depressive effect on soil N2O emissions was also reported by others for digested cattle manure (Rubaek et al. 1996; Petersen 1999; Amon et al. 2006) and pig manure (Vallejo et al. 2006). Nyberg et al. (2004) concluded that some compounds present in AD-manure may have a depressive effect on soil NH3 oxidizers, thereby reducing the supply of substrate for N2O production through nitrification and denitrification. Alternatively, Vallejo et al. (2006) argued that because most easily degradable C present in manure is decomposed during AD, the C remaining in the digested manure is more stable and, therefore, less likely to stimulate denitrification and N2O production as compared with the undigested manure.

The rate of N2O production is primarily dependent on the availability of mineral N (e.g. Granli and Bøckman 1994; Bouwman 1996) and on the microbial characteristics of the soil, some soils promoting denitrification until N2 and others producing a large fraction of N2O (Hénault et al. 1998). The magnitude of the emissions depend on the amount of manure applied, the crop type and the soil temperature and soil moisture content. Application of manure to moisture-retentive soils produces greater N2O emissions than application to free-draining soils (Skiba et al. 1992). Repeated changes in soil moisture status and re-wetting of dry soil promote large, episodic, N2O emissions (Flessa et al. 1995). Application to, or incorporation into, warm soils is also likely to lead to greater emissions than from soils which are cold. However, some studies have shown that the largest N2O emissions occur during thawing of frozen soils (Müller et al. 1997), and the total emissions between November and February were 50% of the total annual flux (Kaiser et al. 1997). Rapid crop growth, and demand for NO3

--N, will reduce N2O emissions by reducing the pool of mineral N available for denitrification (Yamulki et al. 1995). Increased exudation of C from plants may also increase denitrification.

Estimation of nitrous oxide emissions

Current estimates of N2O emissions from soils from N fertiliser applications are based on assumed percentages of the whole waste N content (IPCC 2006). The


24

results above suggest that emissions of N2O are influenced by many waste derived characteristics associated with the waste source and its pre-treatment. Nitrous oxide emissions are also affected by other soil factors such as soil type and time of application in relation to the period of excess rainfall and crop growth. All these factors may impact the short term N2O emissions and therefore differences in waste characteristics may have an important impact on such emission.

Whilst in this study a simple approach based on IPCC 2006 (see Section A. 3.5) recommendations has been adopted, it is recommend that further work is carried out to determine whether waste specific characteristics may be used to produce a more accurate prediction.

A2.4.3 Methane

Soils are the main sink of atmospheric CH4 through the action of CH4-oxidising bacteria (Conrad 1996, Minami and Takata 1997). Ideally agricultural soils should sustain healthy populations of CH4-oxidising bacteria (methylotrophs) that may oxidise CH4 generated in anoxic hotspots in the soil and remove atmospheric CH4.The application of organic wastes to agricultural soil may affect positively or negatively the activity of CH4-oxidising bacteria in several ways. These may include:-

Provision of nutrients – enhancing the growth of methylotrophs. Alternatively however the provision of nutrients may enhance the growth of competitive microbes which out-compete the methylotrophs.

Provision of biodegradable carbon – reducing the soil oxygen content limiting the oxic region in the soil and the growth of the aerobic methylotrophs. This may also increase the amount of CH4 generated in anoxic soil areas, which may increase the methylotrophic activity in aerobic soil areas by supply of soil CH4 as carbon source. If soils are waterlogged and have a high content of biodegradable organic matter methane oxidation may cease and the soils become emitters of CH4 for longer periods.

Change in soil properties - such as moisture content and pH which may positively or negatively affect the survival of the methylotrophs.

Addition of inhibitors of methylotrophs – for example NH3 is an inhibitor of CH4 oxidation and in high concentrations inhibits methylotrophic bacteria. A number of studies have been reported to show that nitrogen addition reduces the uptake of atmospheric CH4 by soil (Conrad 1996).

Ideally the aim should be to enhance rather than reduce the CH4 oxidising capacity of soils as a result of applications of organic matter to soils. Predicting the impact of organic matter applications to soil on CH4 oxidation is difficult and it is recommend that further work is undertaken to evaluate the impact of waste characteristics on this soil function.

A2.4.4 Carbon dioxide

Whilst CO2 is the main greenhouse gas, it is of interest in the context of rate of soil respiration, aerobicity of soil and carbon sequestration in soil. It will be a function of the biodegradability of the organic waste being the product of organic matter aerobic decomposition.


25

A2.4.5 Odours

Odour issues will usually be associated with the waste type and how and when it is spread to land. This is more of a public nuisance than a health issue although adverse reactions do occur in some people (EA 2008). Generally sewage sludge is an odorous material although this is greatly reduced following stabilisation. AD digestates have the potential to be odorous if they contain reduced sulphur species and amines. Generally composted wastes will have fewer potential odour issues.

The assessment is based on an assumption for potential to produce odours based on pre-treatment and its application rate.

A2.4.6 Organic pollutants

Organic pollutants in organic waste recycled to land are of particular concern because of the wide variety of potential organic pollutants available. These may differ significantly in their level of toxicity to different receptors, mobility (leachability) and bioavailability, and persistence (biodegradability) in the pre-treatment processes and soil environment. Many organic pollutants however are biologically active at very low concentrations and therefore may present a risk even when diluted in organic wastes. Several studies have been carried out on organic pollutants in organic wastes focusing on organic compounds such as PCBs, dioxins, PAHs, BTEX, Endocrine disrupting chemicals (EDCs), pharmaceuticals and pesticides. However there is not widespread monitoring of wastes for organic pollutants. The difficulty is that analysis and monitoring for individual organic pollutants is costly..

In the UK the PAS100 compost specification and the draft PAS110 digestate specification only include source-segregated wastes as a precautionary principle in the expectation that such materials have less risk of containing organic pollutants than mixed source organic wastes such as MBT derived compost-like outputs (CLOs) and sewage sludge. Despite this there are potential risks from organic pollutants in such source segregated materials, for example the recent concerns regarding the herbicides aminopyralid and chloropyralid being present in composts (AFOR 2009). Herbicides such as these may be derived from cattle manures where the cattle have been fed grass or silage treated with herbicides and from domestic greenwaste where pesticides may have been used.

In terms of this study it can be assumed that organic pollutants present no agricultural benefits and only pose an environmental risk. The wide variety of chemicals involved suggests that any assessment should be made on case-by-case but this needs further consideration. It is unlikely that sufficient data is available to generalise on different risks from different wastes.

The plant growth test of PAS100 is a screen for the presence of plant toxins and inhibitors that provides some protection for the use of the compost. Similar screening is not carried out for potential animal toxins and the presence of potentially cumulative toxins such as PAHs, PCBs and dioxins. Ecotoxicological methods for assessing organic wastes are available and have been applied, e.g. the soil collembolan Folsomia candida (Domene et al. 2007) but available data is rare and the results indicated that the degree of stability of the organic waste was a major factor. The ecotoxicity was greater in wastes that were less stabilised.

Consideration may be given to simple laboratory screening ecotoxicological tests rather than measuring a long list of chemical pollutants. CEN Technical Committee


26

292/ Working Group 7 (ecotoxicity) has recently published a methodology for the application of ecotoxicity tests for wastes and a list of recommended tests to fulfil both hazard assessment and environmental risk assessment functions .

A2.5 Characteristics by waste type

The relevant waste characteristics may be in some part a feature of specific waste groups, although the approach taken here is to assume that there is considerable overlap between waste groups and that focus on the characteristics of the waste under scrutiny would be more useful. The following provides a brief discussion of the main waste groups.

A2.5.1 Animal manures

Animal manures are principally derived from cattle, poultry, pigs and sheep and come in the form of either liquid slurries or more solid manures. These materials consist of animal faeces and urine mixed with bedding material (typically straw). They are usually rich in readily available N from the urine and have been applied to agricultural land principally as fertiliser to supply N, P and K. Guidance on the nutrient content and how to apply animal manures is given in RB209.

The majority of organic matter in manures and slurries is not considered to be highly biodegradable, as it comprises material that has already been through a decomposition process in the animal gut. However, a proportion of the carbon in manures is in the form of volatile solids (VS) and these compounds are precursors of emissions of methane (CH4) and non-methane volatile organic compounds (NMVOCs) and may also be a source of readily metabolizable carbon for soil microbes. Manures are increasingly being considered as a feedstock for anaerobic digestion which would be expected to impart some changes in their characteristics, in particular the reduction in the proportion of C present due to microbial breakdown of the organic matter to biogas (CO2 + CH4).

A2.5.2 Sewage sludge

Sewage sludge (often called biosolids) is derived from sewage treatment. Several forms of treatment may have been applied giving rise to sewage sludge of differing characteristics. Typically sewage sludge may be anaerobically digested in a wet digestion process and then either used as a liquid sludge or dried to produce a solid cake. Sludge cakes may be prepared by dewatering by mechanical presses and centrifuges, which means that some of the readily available nutrients such as ammonium (NH4

+) and K may be lost. Cakes may also be prepared by drying with heat which may also mean that some losses of nutrients such as NH3 are lost.

Another common sewage sludge treatment is lime stabilisation where the sludge is mixed with lime to increase the alkalinity of the sludge and at the same time sanitise the sludge by the additional heat generated. Guidance on the fertiliser value and use of sewage sludge in agriculture is also given in RB209.


27

A2.5.3 Anaerobically digested wastes

Anaerobic digestion is favoured as a biological treatment process for organic wastes as it allows recovery of energy from the biogas produced. Anaerobic digestion does not release large amounts of energy for microbial growth and therefore in general the organic matter is decomposed to a mixture of CO2 and CH4 with little production of microbial biomass. Not all the organic matter is decomposed and typically the lignin material is not degraded under anaerobic conditions.

Economic anaerobic digestion is carried out for a short period of about 15 – 25 days and therefore complete stabilisation of the organic material does not occur. Digestion, by mineralization of organic matter may result in a digestate rich in nutrients h, especially NH4

+, (increasing the amount of RAN).

Anaerobic digestion may be carried out at low solids concentrations (up to 10% dry matter content) in a wet digestion producing a slurry product that may have similar characteristics to manure slurries. Anaerobic digestion may also be carried out at higher solids contents in dry AD processes. The solid whole digestate will similarly be enriched in readily available NH4

+ and may in the first instance be considered as having similar fertiliser properties as solid manures. Digestates may be mechanically de-watered and/or heat dried. The raw digestate, the liquid residue from de-watering and the de-watered solid may be recycled to soil. The characteristics of these products may vary in their fertiliser value and environmental risks.

Organic wastes anaerobically digested are typically the more putrescible material such as food and animal wastes. Often these may be source segregated and may in future be monitored by the draft PAS110 digestate specification and regarded as products if their production complies with the Quality Protocol for digestates (EA 2009). Many MBT systems also include anaerobic digestion of the mixed waste putrescible fraction and may also be applied to land, although not as products as defined in the PAS 100 Quality Protocol.

A2.5.4 Composts

Composting of organic wastes is an established approach to produce a compost product for use on land. Composting is carried out under aerobic conditions for an extended time period resulting in significant decomposition of the organic waste. Therefore the resulting matter is usually resistant to significant further microbial decomposition.

Composting is typically carried out on material with a moderate biodegradability such as greenwaste so that the biological oxygen demand may be met and the composting process remains aerobic. Where highly biodegradable materials are composted they are usually mixed with bulking agents to dilute the oxygen demand and provide an open structure for aeration. During composting a significant proportion of the organic matter is converted to microbial biomass, part of which then may itself decompose. Therefore much of the end material is derived from microbial biomass. The lignin is also significantly degraded during composting. Consequently the nature of the organic matter in compost may be different to manures, sewage sludge and anaerobic digestates.

Composting of biosolids is seen as a process that stabilises the organic matter content and reduces the volume and mass of the waste. Application of such


28

composts to land has potential to recover value and avoid disposal to landfill (Courtney and Mullen 2008).

Maturity, in the context of composts, is related to the effect of composts on crops and indicates the presence or absence of phytotoxins. Therefore, immature or phytotoxic composts may have a negative impact on plant growth. Stability is a stage in the decomposition of the organic matter (Epstein 1997) and is, therefore, a term related to the microbial decomposition or microbial respiration activity of the composted matter.

Bernal et al. (1998) found that the ratio of water-extractable C to organic N (Cw/Norg) correlated well with most of the other maturity indicators, rendering it the most suitable for evaluating compost maturity. A definition of 'compost stability' is included in a working document by the European Commission (2001), since there is no limit for stable composts set by the EU. According to that definition, stable composts should have, among others, respiration activities less than 10 g O2/kg dry matter after four days (AT4 test).

Biosolids tend to be slightly acid, whereas after composting they tend to be neutral. The heat generated during composting also kills pathogens and reduces the risk of these in compost products. Esteller et al. 2009 found total and faecal coliforms were 2.4 * 108 and 9.3 * 107 MPN g-1 (dry basis) in biosolids, whereas in the composted biosolids, total and faecal coliforms were 1500 MPN g-1 and 30 MPN g-1: a 99.9% decrease.

Komilis and Tziouvaras (2009) concluded that phytotoxicity test germination indices are highly dependent on the type of seed used. A compost that is phytotoxic to a certain seed can enhance the growth of another seed. Therefore, the use of germination bioassay tests may be unreliable when it is used to assess compost maturity, since a common threshold for germination index (GI) to indicate maturity cannot be established for all types of composts. A stability test, alone, is not enough to ensure high compost quality. For example, poultry manure compost was highly phytotoxic to most seeds, and, therefore, immature, but resulted in a low respiration activity, which is indicative of stable composts. Therefore, it appears that determining compost quality requires a simultaneous use of maturity and stability tests.

A2.5.5 Paper pulp wastes

Paper pulp mill sludges may supply large amounts of organic matter and are not generally contaminated by heavy metals and organic pollutants. As such they are potentially useful soil amendments with potential to improve soil physical properties and recycle carbon to soil (Calace et al. 2005; Carpenter and Fernández 2000; Gagnon et al. 2001). Nunes et al. (2008) cite several papers that have reported substantial benefits to soil fertility and crop yields from application of paper sludge wastes to low-carbon soils (Levy and Taylor 2003). These sludges have been claimed to significantly increase SOM (Rotenberg et al. 2005), provide a range of nutrients (Feldkinchner et al. 2003), increase soil pH (Battaglia et al. 2007), increase beneficial soil organisms and reduce plant pathogens, increase water and nutrient retention (Foley and Cooperband 2002). There are also claims of improving the soil‟s ability to suppress crop diseases (Madrid et al. 2007; Rotenberg et al. 2005).

However paper sludge may have C:N ratios of 100–300:1, and land applications of sludges with large C:N ratios may lead to a temporary immobilization of soil N and


29

require addition of N to compensate. Conversely, large amounts of N applied with paper mill sludge could potentially cause nitrate leaching (Feldkinchner et al. 2003).

Nune et al. (2008) evaluated a paper sludge waste with C:N and C:P ratios of 11 and 19 respectively. Citing critical C:N ratio of 20–30 (Cordovil et al. 2007) and C:P ratio of 40–50 (Fageria et al. 2007) Nune et al. (2008) forecast mobilization of N and P when this sludge was added to the soils of the trial. Sludge applications increased SOC, but only at the largest applications (>80 t/ha for the sandy soil and > 120 t/ha for the heavier soil). The N content of the soils generally increased with increasing sludge application but very large applications have been shown to immobilize N. Soil P increased with increasing sludge application, however, in contrast with N which increased in grain in proportion to the sludge applications, there were no increases in grain P from sludge application. This was attributed to resorption or precipitation of soluble P released from SPS by Ca compounds.


30

A3. METHODOLOGY FOR ASSESSMENT OF AGRONOMIC BENEFIT

A3.1 Overview

An assessment has been carried out for agronomic benefits and environmental impacts by normalising the application rates of different wastes to soil under two conditions.

Use as a fertiliser by normalisation to the same N fertiliser value.

Use as a soil conditioner by normalisation to give the same increase in soil organic carbon.

Following normalisation of waste application rates the loadings of other parameters have then been estimated and compared. The comparison principally examines the difference between the impact of the waste in question and that of a typical sewage sludge applied to soil. Sewage sludges are an acceptable organic soil amendment and provides a benchmark for comparison of other wastes. The comparison has used mean values of the different waste groups and treatments (Tables 10.1 and 10.2 in Appendix 10). The approach may be applied to any specific waste composition to assess the benefits and disbenefits, e.g. the maximum or particular percentile value.

The following sections describe the methodologies and assumptions applied in the normalisation of loads and the comparison of agronomic benefit. This methodology has also been used to compare some of the potential environmental impacts especially gaseous emissions and soil quality.

A3.2 Calculation of loadings by normalisation to N fertiliser value

In RB209 guidance for manure applications as fertiliser is based on estimation of the readily available nitrogen (RAN) and the corresponding crop available nitrogen (CAN) for the particular manure and soil. For example the RAN for pig slurry is estimated as between 40-60% of the total N. But if applied to a sandy or shallow soil in the autumn it is estimated that the CAN is only 5% of the total N. The difference is then a potential source of environmental N emissions to air and water. This guidance does not fully consider the mineralisable ON of the manure as much of the TN is already present as RAN. For many wastes the content of RAN may be small and most potential N nutrient value may be in the form of ON mineralised during decomposition of the waste in the soil. For such wastes the fertiliser value may be better based on potentially available nitrogen (PAN) which is the sum of RAN and the fraction of ON that is mineralised during decomposition.

Of the waste N parameters, PAN, RAN and CAN may all be potentially used to normalise organic waste loadings. In this study PAN estimations have been used, as this parameter may be applied to all wastes and not limited to those with a high RAN content.


31

The waste loadings have been normalised to provide a 100 kg PAN/ha which from RB209 represents the N requirement of an autumn and early winter wheat for a light sandy soil with SNS index of 1.

Waste PAN content estimations based on actual measurements are expected to be rare as this involves measurement of N released during decomposition of the waste. Higgins et al. (2005) described the calculation of PAN from the total N and RAN where PAN = RAN + (TN-RAN)/2. This does not take into account the biodegradability of the organic N of the decomposing organic matter, and whilst it may be applicable for some raw wastes it would not be appropriate for stabilised composts where most of the N is organic N and poorly biodegradable.

Another factor for N mineralization is the C/N ratio of the decomposing organic matter. If the C/N ratio is high (> 30) then a significant fraction of the N will be re-used as N source for microbial growth and in this case although the organic matter C will decompose and N mineralization will not be significant. This is a known feature for wastes such as high C/N ratio paper sludges. PAN estimations for organic materials with low RAN contents clearly require some assumptions and estimations that will have some errors.

A3.2.1 Method to estimate mineralisable nitrogen from the organic nitrogen

For this estimation it has been assumed that the organic nitrogen mineralisation is in direct proportion to the extent of C-mineralisation in the short term (first year) in the soil. Waste aerobic C-mineralisation studies of this length of time are rarely carried out. Therefore, in most cases it is assumed a C-mineralisation value based on expert knowledge of the waste and the pre-treatment it has received, and short term aerobic and longer term anaerobic biodegradability tests (e.g. DR4 and BM100 tests, Godley et al. 2007) carried out on such materials.

The percentage ON-mineralisation is then amended by a factor defined by the C/ON ratio of the waste. These factors are a guess-estimate based on expert knowledge of C/ON ratios but can be modified when additional evidence is available.

Therefore the percentage ON-mineralisation from the organic N (%ON-min) is given by:

%C-min x FC/ON where FC/ON is the C/ON ratio reduction factor given in Table 3.1.

Table 3-1 Estimation of organic nitrogen mineralisation factor from waste C/ON ratio

Waste C/ON ratio <10 11 - 20 21 - 30 31 - 40 41 - 50 > 51

C/ON reduction factor FC/ON 1 0.8 0.6 0.4 0.2 0

The estimation of PAN for any waste is then given by

PAN = RAN + (ON x %ON-min/100) Equation 1

Example calculation to normalise organic waste loadings to PAN


32

The target PAN is set at 100 kg N/ha as fertiliser. This is the guideline value given in RB209 for a light sand soil with SNS index of 1 for the growth of autumn and early winter sown wheat.

A mature compost is available with characteristics given in Table 3.2

Table 3-2 Characteristics of mature compost for example PAN calculation

Total C g/kg DM

Total N g/kg DM

RAN g/kg DM

ON g/kg DM

C/ON ratio

Biodegradability in soil (%C-min)

200 10 0.5 9.5 22.2 10

From Table 3.1 the FC/ON for a waste with C/ON of 22.2 = 0.6

Consequently the %ON-min = 10 x 0.6 = 6% where 10 is the percentage biodegradability of the carbon

PAN is then given by equation1 as

DMtkgkgDMgPAN /07.1/07.1100

65.95.0

The PAN is then 10.7% of the total N. This gives a reasonable estimate for a mature compost.

The compost loading required to give the normalised loading of 100 kg PAN/ha

= 100/1.07 = 93.5 t DM/ha.

A3.3 Calculation of loadings of wastes as soil conditioner

In this scenario the organic waste is applied as a soil conditioner to increase the organic matter content of the soil. The different organic wastes will have different biodegradabilities depending on their source and pre-treatment. It is assumed that a significant amount of the residual biodegradability of the waste would be lost in the first year of application and that the remaining waste derived organic carbon would be biostabilised and considered to contribute to the soil organic carbon (SOC) content.

Organic waste loads to the soil may then be based on normalising the residual waste derived organic carbon in the soil after one year. In this way more unstabilised wastes would need to be applied to provide the same increase in soil organic carbon after one year compared with a stabilised waste. In most soil studies the waste decomposition is virtually complete within the first year and hence the same waste %C mineralisation assumptions may be used.

Therefore the stable waste carbon is given by:


33

100

min%1

CC

The plough depth for agricultural soils is taken as about 25 cm and the dry bulk density of the soil assumed as 1.2 kg l-1. This gives an approximate soil DM content of 3,000 t DM/ha. A target SOC increase for wastes applied as soil conditioners may then be set for the normalisation of loads. For poor sandy soils SOC contents are about 1% (10 kg C/t DM, 30 t C/ha). Doubling this SOC content by addition of 10 kg C/t soil DM from the organic wastes would seem a reasonable approach for normalisation purposes and equates to the addition of 30,000 kg OC/ha. An increase in SOC of this scale may be expected to significantly improve soil structure.

Many organic wastes (particularly mature composts) contain a significant amount of inorganic matter which will add to the inorganic mass of the soil and therefore dilute the SOC concentration. An adjustment for this is possible but increases the complexity of the calculation and would mean comparison of impacts would be more difficult as different loads would be applied to different soils. Consequently, for this study, this adjustment has not been carried out.

Maintaining SOC concentrations may be achieved by repeated addition of small amounts of organic wastes. This could be an alternative scenario which may be worth considering (see for example Barral et al. 2009). The approach used in this study represents a one-off high application of waste as a worst case scenario for impact assessment comparison.

A3.3.1 Example calculation of loading as soil conditioner

Taking the above compost (Table 3.2) the C mineralised during the first year in the soil is estimated as 10%. Therefore the stable OC in the organic waste is given by TOC x (1- 10/100) = 180 kg/t of waste which then equates to an application of 30,000/180 = 166 t DM/ha.

In this study the loadings have again been compared with sewage sludge. However, this is not a material that would typically be applied to significantly increase SOC as it is much more biodegradable than, for example, mature compost. However the sewage sludge benchmark provides a useful point of comparison with organic wastes used as both N fertiliser source and soil conditioner.

A3.4 Comparison of ammonia emissions from loadings

Agriculture represents 80% of ammonia emissions in the UK, principally derived from the high NH3 content of animal manures. Emissions are mainly from livestock housing, manure storage and the spreading of the manures to land (Defra 2002). Ammonia emissions represent a loss of beneficial N-fertiliser in the recycled manure as well as the environmental impacts following the deposition of the NH3 from the air. Agricultural NH3 emissions have been extensively studied and general principles of manure management are available to minimise the emissions when applying manures to soil.

The source of NH3 emission from livestock manures following application to land is the total ammoniacal-N (TAN) (Webb and Misselbrook 2004 and references cited therein). Hence studies reporting NH3 emission following application of manures to


34

land usually express those emissions as a % of TAN applied in the manure. The total amount of NH3 emissions from livestock manures during and after field application depends on:

properties of the manure, including viscosity, TAN content, C content and pH;

soil properties such as pH, cation exchange capacity, calcium content, water content, buffer capacity and porosity;

meteorological conditions including precipitation, solar radiation, temperature, humidity and wind speed;

the method and rate of application of livestock manures, including, for arable land, the time between application and incorporation, and method of incorporation;

the height and density of any crop present.

Anaerobic digestion results in the mineralization of N as NH3 which remains within the digestate during the AD process rather than being lost as occurs during aerobic composting. Therefore AD digestates represent a potentially greater risk of NH3 emissions than composts. However, the amount of NH3 in AD digestates and the potential impact may vary depending on the type of AD process and how the digestate is subsequently treated after the digestion process (i.e. whether the digestate is used whole or is dewatered and/or heat dried, and whether the liquid from dewatering is also applied to land). In comparison with manures, digestates are likely to be more alkaline. Consequently more of the mineral N is present as free NH3 and more readily volatilised.

It is interesting in this context to note that anaerobic digestion of manures is likely to increase the amount of NH3 and the pH content of the manure. This would increase risks of NH3 emissions from digested manures compared with raw manures. However, the reduced dry matter and viscosity of the digestate allows more rapid infiltration into soil potentially counter-balancing the greater NH3 and increased pH content. Few studies have reported comparison of emissions from raw and digested slurry, those that have did not indicate a significant difference in emissions of NH3 (discussed in Section A2). For the purposes of this study it is proposed that potential NH3 emissions during applications of wastes may be estimated on the basis of their:

TAN concentration;

pH; and

% dry matter (DM).

Table 3.3 below, taken from the Inventory of Ammonia Emissions from UK Agriculture (IAEUKA 2007) indicates a broad relationship between the % DM of livestock slurries and typical emissions of NH3, as a % of TAN (emission factor EF), when those manures are applied to land by surface application. The EF for FYM and poultry manures is independent of %DM, but the means are cited below for information (Defra project NT2006).


35

Table 3-3 Example estimation of ammonia emission factor (EF) as percentage of total ammoniacal nitrogen (TAN)

Manure type % DM pH EF % TAN Modifier for dry matter

Cattle slurry 7.8 7.4 47.5 EF1 = EF x ((12.3 x DM)+50.8)/100

Cattle FYM 23.0 8.0 68.3 N/A

Pig slurry 3.7 7.7 24.6 EF1 = EF x ((12.3 x DM)+50.8)/100

Pig FYM 26.0 7.9 68.3 N/A

Layer manure 35.0 7.8 52.3 N/A

Poultry litter 60.0 8.2 52.3 N/A

This table is a simplified version of the approach used in the IAEUKA for cattle slurries for which the emission factors (EF) are also modified for time of year (wet and dry soil conditions) and land use (arable and grassland).

For this study the same formula:

EF1 = EF x ((12.3 x DM)+50.8)/100

has been used for the estimation of NH3 emissions but used the RAN value instead on TAN as in most cases RAN data is based on NH3 determinations only and rarely includes nitrate. Also an upper limit of 50% NH3 emissions has been set as a realistic maximum if the equation gives a greater value.

A3.5 Nitrous oxide emission estimation

Soils are a major source of atmospheric N2O which is a potent greenhouse gas and is derived from the microbial metabolism of N in soils. Emission of nitrous oxide (N2O) from soil following application of organic manures to land is both more complex and long-term but also a much smaller proportion of the total N applied in manure than is emission of NH3. In soil, two microbial processes produce N2O predominantly:

nitrification, i.e. the oxidation of ammonium (NH4+) to NO3

- and

denitrification, i.e. the reduction of NO3- to gaseous forms of N, ultimately N2O and

N2.

It is not only the TAN in the manure at the time of application that is nitrified, but also NH4

+ which is produced by mineralization of organic-manure in the manure. Hence emissions of N2O are expressed as a % of total- manure N.

When estimating the potential for N2O emissions IPCC (2006) uses an EF of 1% of total N applied for all organic manures. Hence it is proposed to use the IPCC default factor for N2O emissions following applications of organic manures to land, when estimating emissions following application of wastes to soil.


36

A3.6 Carbon dioxide emission

The decomposition of organic matter in soils and the potential impact of waste composition and waste treatment were discussed in Section A2.3.10. For this study a simple view of estimating the extent of carbon degradation of the applied waste within the soil has been taken, based on assumed biodegradability characteristics of the waste. The waste organic carbon decomposed is assumed to be emitted to air as CO2.This provides a simple initial approach for estimating and comparing CO2 emissions from the different wastes, although this approach may be improved with more complex models when specific waste data becomes available.

A3.7 Methane emissions

Soils are an important sink of atmospheric CH4 through the activities of CH4 oxidising bacteria in soil. The application of organic wastes may influence this important soil activity with possible scenarios being inhibition of the soils CH4 oxidising capacity and turning soils into CH4 emitters through methanogenic activity in anoxic zones of the soil (see Section A2.4.3). This is a complex topic which has not been addressed in this study and therefore evaluation of impact on CH4 emissions has not been included.

A3.8 Phosphorus nutrient value assessment

The earlier discussion (Section A2.3.5) indicates the uncertainty with estimating available P from organic matter applications to soil. The available P may not be reliably measured by mild extraction methods such as Olsen-P and that stronger methods may be more applicable. For this assessment a P availability of 60% has been assumed for all the wastes as a conservative value. The assessment of P nutrient addition can be made by reference to the crop P requirement. For this, the P requirement for autumn and early winter wheat sown on a light sand soil with a P index of 1 and where straw is removed has been used. The amount of P required is 95 kg P2O5/ha (41.5 kg/ha)

A3.9 K, Mg and S nutrient value assessment

The discussion in Section A2.3.6 to 8 indicate there are fewer concerns from adding excess K, Mg and S to soils compared with other elements present in the waste.

The assessment is simplified to a comparison between wastes, where for Mg and S the assessment compares total values, but for K assumes a 90% availability of the total K as applied in RB209.

The beneficial K may be assessed by comparison with the K required for autumn and early winter wheat sown on a light sand soil with a K index of 1 and where straw is removed the amount of K required is 95 kg K2O/ha.

Magnesium and S are only typically required when a soil deficiency for these nutrients are detected. In this case the recommended applications are up to 50 kg Mg/ha and 16 kg S/ha.


37

A3.10 Assessment of metal applications

Virtually all organic wastes will contain potentially toxic elements (PTEs) and the most commonly monitored of these are the metals required by the 1996 DoE Code of Practice for the Agricultural Use of Sewage Sludge – cadmium (Cd), chromium (Cr), copper (Cu), lead (Pb), mercury (Hg), nickel (Ni) and zinc (Zn). Whilst some of these are trace elements (Cu, Ni, Zn) required for growth it is rare for these to be limiting and most concerns would be for their excessive application to soils.

This assessment compares the increases in total soil metals as a result of the addition of organic wastes to soil. This has been carried out by calculating the number of normalised applications needed to exceed the soil metal limits for a typical soil. The assessment also estimates the mean 10 year average metal application rates and compares these with the metal application rates limits in the UK Code of Practice for sewage sludge applications to soil (Defra 2003).

A3.11 Assessment of organic pollutants

Data for organic pollutants in organic wastes applied to soil are very limited. However an approach similar to that used for metals may be made as a means of assessing impact. For this approach, limit values of organic pollutants in agricultural soils need to be defined. Then the number of waste applications required to exceed these limits may be estimated for a typical soil (with an assumed initial level of organic pollutant present).

The soil limits that may be used need to be developed through a risk assessment approach for example, as used by the Environment Agency for assessing contaminated land by the CLEA model (EA 2009b), although soil guideline values are not currently available for many key organic parameters.

This approach has been demonstrated here for a limited number of pollutants and may be developed further if considered a useful methodology. The approach assumes there is no microbial degradation of the organic pollutant in the soil and so can be considered a conservative assessment.

A3.12 Assessment of pathogens

There is insufficient data to assess the pathogens in the organic wastes. As pathogens are living organisms with the ability to grow and reproduce it is difficult at this stage to define or compare impacts from different wastes. Pathogens potentially associated with different wastes are discussed in Section A8.


38

A4. ASSESSMENT AGRONOMIC VALUE OF WASTES AS FERTILISER

This section describes the results from assessing the agronomic value of the wastes. The comparison is based on normalising application rates to provide 100 kg PAN/ha as nitrogen fertiliser. The assessment was carried out for different wastes and for different treatments using the mean composition shown in Annex A to this report. Sufficient data was not available for every waste from Table 2.2 and 2.3 to be considered. Where an assessment could be made using an assumed value this data is noted in the tables.

A4.1 Total dry matter loading comparison

A4.1.1 Total dry matter loading for different wastes

The estimation of the waste DM loading (Figure 4.1) required to achieve an effective PAN of 100 kg/ha is very dependent on the waste characteristics with respect to its mineral N content, and the decomposition of the organic matter in soil, which may either mineralise organic N (contribute to PAN) or consume mineral nitrogen (subtract from PAN). See Section A2 for full list of wastes.

0.0

5.0

10.0

15.0

20.0

25.0

30.0

35.0

40.0

45.0

50.0

26 11 12 13 14 15 16 17 18 19 23 26 31 32 33 34 35 36 37 3840

.140

.440

.5 41 42 44 45 46 47 48 4950

.650

.7 51 52 53 54 55 71 73 74 75 76 78 N

Waste type code

Ap

pli

cati

on

to

PA

N (

t D

M/h

a)

120 55 121 487148

Figure 4.1 Dry matter loadings (t/ha) for different wastes to achieve PAN of 100 kg N/ha) (mauve bar - sewage sludge standard where PAN is 50% TN)

This graph indicates that for many wastes the loadings required to give comparable PAN concentration may be similar to the loadings of sewage sludge (e.g. most


39

manures). There seems to be much more variation between the green waste and food wastes than the manures which may reflect the greater diversity of sources of some of these wastes. Some waste input types (notably wood, de-inking sludges from paper recycling and dredgings) would not normally be considered as sources of N fertiliser and these are shown here to require the highest loadings. The data analysis indicates that chipboard and MDF may be suitable but the TN determinations used in the data analysis are considered to be unusually high and may be due to errors in the original data, although these were as reported. The mechanical biological treatment residues (most of which are composted) gave a high loading and therefore are not considered to be good material as N fertilizer. Wastes that require significantly higher loadings than sewage sludge to give the same PAN benefit may not be suitable for land application as they may carry excessive quantities of other materials as contaminants.

This visual comparison of loadings by waste type could aid future decision making, and could be used as a screening step for materials applied to land in the future. This would allow focus on those wastes that may not provide a good N source and allow consideration of whether they contribute other benefits or harm.

A4.1.2 Total dry matter loadings by treatments

For this analysis it has been assumed that the composts are fully stabilised and consequently there is little N made available during decomposition. Loadings to PAN by treatment type on this basis indicate (Figure 4.2) that application rates are much higher for the composts compared with sewage sludge by about a factor of 10 (codes CL, CM and C). These higher loading rates would be expected of composted material and is further indication that this comparison approach reflects reality and may lead to focus on whether composts are good as fertilizers for other plant nutrients other than N.

0.0

10.0

20.0

30.0

40.0

50.0

60.0

26 CL CM C DM D H LS L M N Q RL R TM T U

Waste treatment code

Lo

ad

ing

to

10

0 k

g N

/ha

PA

N (

tDM

/ha

)

Figure 4.2 Loadings of waste treatments to achieve the PAN loading of 100 kg N/ha compared with benchmark sewage sludge (purple bar)


40

A4.2 Total nitrogen loadings

The loadings to PAN for the different wastes and waste treatment would apply very different amounts of TN to the soil. Estimations of TN applications indicate that for most wastes (Figure 4.3) the TN loadings are similar to the benchmark sewage sludge. However many application rates approach NVZ limits of 250 mg/ha on non-grassland, indicating that the application of some wastes as N fertilisers would be limited in some areas. Notable wastes that have greatly exceed the NVZ limit of 250 kg N/ha include sludges from biological treatment of industrial waste (code 23), wood waste (code 37), ABPR wastes (code 40.4) and MBT residues (code 54).

The analysis for the waste treatments (Figure 4.4) indicate that the total N for the compost exceed the NVZ zone limits by a factor of 4. Clearly this demonstrates that composts may not be applicable for use as N fertilisers and that application rates would need to be significantly lower to comply with NVZ limits and for their corresponding benefits to of demonstrable value at these lower application rates.

0

100

200

300

400

500

600

700

800

900

26 11 12 13 14 15 16 17 18 19 23 26 31 32 33 34 35 36 37 3840

.140

.440

.5 41 42 44 45 46 47 48 4950

.650

.7 51 52 53 54 55 71 73 74 75 76 78 N

Waste type code

To

tal N

ap

plic

ati

on

(k

g N

/ha

) fo

r lo

ad

ing

s t

o P

AN

of

10

0 k

g N

/ha

Figure 4.3 Total N loadings for waste types applied to achieve a PAN of 100 kg N/ha


41

0

200

400

600

800

1000

1200


Waste treatment code

To

tal N

lo

ad

ing

s (

kg

N/h

a)

fo

r lo

ad

ing

s t

o

PA

N o

f 1

00

kg

/ha

Figure 4.4 Total N loadings for waste treatments applied to achieve PAN of 100 kg N/ha

A4.3 Phosphorus loadings

As detailed in Section 3, for this assessment a P availability of 60% has been assumed for all the wastes as a conservative value. The assessment of P nutrient addition can be made by reference to the crop N addition. Having supplied the crop N requirement the extent of the crop P requirement met by the waste can be compared using the same crop table. For autumn and early winter wheat sown on a light sand soil with a P index of 1 and where straw is removed the amount of P required is 95 kg P2O5/ha (41.5 kg P/ha).

The results for the available P loadings for the different waste types are shown in Figure 4.5. Sewage sludge is usually recognised as being P rich and sewage applications for N fertilizer value can therefore potentially provide excessive P. The sewage sludge mean composition gives an available P loading of about 67 kg P/ha (total P loading of 112 kg/ha) when applied at the PAN application rate of 100 kg N/ha. This has exceeded the benchmark value of 41.5 kg P/ha which confirms that caution is required for P loadings from sewage sludge and that the actual value of the percentage availability of the P a critical parameter.

Most other wastes have lower P loadings and may be of value as N and P fertilizers even if their total DM loadings exceed that of sewage sludge and assuming there are no other undesirable contaminants carried with wastes. This study indicates that some livestock manures (cattle slurry, chicken manure), and wastes from animal products (gut contents, ABPR wastes), may carry a high risk of applying excessive available P. The differences between waste types present options to select appropriate wastes in combination to optimise the N and P fertilizer requirements of any soil.


42

Figure 4.5 Available P loadings from application of waste types to a PAN of 100 kg N/ha

Calculated available P loadings for composts (Figure 4.6) applied as N sources to PAN were similar or slightly higher the sewage sludge loadings despite the composts having lower P contents. This is a consequence of the higher DM loadings required to achieve PAN relative to sewage sludge. This would mean that the available P would potentially exceed the plant requirements by a greater margin and pose a greater risk of adverse environmental impact. However, applying compost at the required P loading in combination with another source of N would be a means of utilising composts as effective P fertilisers.

A key consideration in this evaluation would then be the availability of the P in the organic waste. There is no data available that can be applied to differentiate the available P from the total P.


43

Figure 4.6 Available P loadings for waste treatments applied to a PAN of 100 kg N/ha

A4.4 Loadings of other major nutrient

Organic wastes would provide other major plant nutrients (K, Mg and S) as well as N and P. Whilst these present lower potential for environmental harm than N and P it would be important to understand how much is added to soil in order to optimise organic waste applications to land.

A4.4.1 K loadings

The beneficial effect of waste K contents may be assessed by comparison with the K required for autumn and early winter wheat sown on a light sand soil with a K index of 1 and where straw is removed the amount of K required is 95 kg K2O/ha (79 kg/ha). Figures 4.7 and 4.8 shows the estimated available K loadings (90% of total K) for the waste types and waste treatments compared with sewage sludge.


44

Figure 4.7 Available K loadings for different wastes types to give PAN of 100 kg N/

Figure 4.8 Available K loadings for different waste treatments applied to give a PAN of 100 kg N/ha

These results indicate that most wastes and especially the composted wastes supply much more K compared with sewage sludge. The sewage sludge application of 28


45

kg K/ha is much less than the benchmark crop requirement application of 79 kg K/ha which confirms the generally accepted view that sewage sludge is K poor. However the K applications would in some cases greatly exceed the benchmark K loading for crop growth. Whilst this may not be considered an environmental issue of concern, it may be advisable to consider any impacts further.

A4.4.2 Magnesium loadings

Figures 4.9 and 4.10 show the loadings of total Mg for the different waste types and waste treatments when applied to give a PAN of 100 kg/ha.

For Mg there is no data for several of the waste types and waste treatments. However, Figures 4.9 and 4.10 indicate also that for most wastes and especially for compost the Mg loadings would exceed the sewage sludge applications and the benchmak crop requirement of 50 kg Mg/ha. The results suggest that some waste would have particularly high loadings, i.e. sludges from biological treatments (314 kg/ha), de-inking sludges (328 kg/ha), dredgings (334 kg/ha) and construction and demolition wastes (7880 kg/ha).

The results indicate that many wastes would be very good sources of Mg and that there is some potential risk of supplying excessive Mg for typical application rates. The environmental impact of this is not considered a high risk, although it is advised that this is considered further.

Figure 4.9 Total Mg loadings for different waste types applied to a PAN of 100 kg N/ha


46

Figure 4.10 Total Mg loadings for different waste treatments applied to PAN of 100 kg N/ha

A4.4.3 Sulphur loadings

The S loadings for the waste types and waste treatments are shown in Figures 4.11 and 4.12. Most waste types have a similar S loading compared with sewage sludge when applied to a PAN of 100 kg N/ha and the crop benchmark requirement of 16 kg S/ha. High loadings were estimated from sugar processing (46), De-inking sludges (71) and dredgings (73). The environmental impact of high S loadings to soil are not thought to be significant although the implications of high S in anoxic soils may be of more concern from production of toxic sulphides and odours. Dredgings would not normally be applied as a fertiliser but application rates are permitted at high rates as it is considered to be a soil like material. De-inking sludges may be considered biodegradable and may load a high initial BOD to the soil and create anoxic conditions. This may be an area for further investigation.


47

Figure 4.11 Total S loadings for waste types applied to achieve PAN of 100 kg N/ha

Treatment type

The available data for waste treatments is incomplete with few data points. However the results indicate that composts S loadings would be higher than sewage sludge and consequently exceed the crop S requirements. As composts would not be expected to have a high BOD there would be a low risk of anoxic conditions from composts. The impact of excessive S addition on other environmental compartments may need some investigation to provide re-assurance that there are no high risk issues.


48

Figure 4.12 Total S loadings for waste treatments applied to PAN of 100 kg N/ha

A4.4.4 Summary major nutrient loadings

In general, for most of the waste types, the nutrient loadings of the major elements K, Mg and S are similar to or exceed the loading from sewage sludge when applied at the normalised PAN loading of 100 kg N/ha. This indicates that most organic wastes would be good sources of these major plant nutrients. The sewage sludge applications for K, Mg and S were 28, 14 and 44 kg element/ha respectively with the K and Mg matching plant requirements and the S exceeding plant requirements. High applications of K and Mg are not considered a high environmental risk. Application of S requires further consideration especially if associated with a high waste BOD. If it is present in reduced form, it may cause toxicity problems, odours and soil acidification when oxidised to sulphate by microorganisms.

At PAN loadings of 100 kg N/ha composts had much higher contents of K, Mg and S than sewage sludge, and could be potentially used as fertilisers for these elements at lower loadings than used in this comparison. This is an indication that different organic wastes may have different nutritional benefits and that consideration of benefits is multi-dimensional. Composts clearly supply crop nutrients other than N.

One of the reasons that sewage sludge is low in K is that this element is very soluble and would be partitioned mainly in the waste water treatment plant effluent rather than the solid waste. Where there is little leachate formation as during most composting processes the, levels of K would be expected to be higher. A simple comparison of the P, K, Mg and S contents of sewage sludge (code 26) and composts (code C) demonstrates these differences in concentrations of plant nutrients (Table 4.1).

Table 4-1 Comparison nutrient content in sewage sludge and composts (g/kg DM)

Waste Total N Total P Total K Total Mg Total S

Sewage sludge (26)

38.7 21.7 5.4 2.8 8.5

Compost (C) 20.2 3.6 8.5 2.8 7.8

Digestates from wet AD processes treating food wastes and similar materials may be expected to have similar nutrient levels to sewage sludge if there is mechanical dewatering. Digestates from dry AD processes where there is less leachate generated may retain more plant nutrients. This difference in AD process especially with respect to the amount of water addition and mechanical de-watering may mean that AD offers opportunities to produce digestate of different fertiliser value. Commercial data of digestate from the organic fraction of MSW where the digestate had undergone extensive washing via a wet AD process was obtained for this study. This digestate had a remarkably low metals content compared with other MSW derived CLOs.


49

A4.5 Metal loadings from application of wastes to a PAN of 100 kg N/ha

Virtually all organic wastes will contain potentially toxic elements (PTEs) and the most commonly monitored of these are the metals required by the 1996 DoE Code of Practice for the Agricultural Use of Sewage Sludge – cadmium (Cd), chromium (Cr), copper (Cu), lead (Pb), mercury (Hg), nickel (Ni) and zinc (Zn). Whilst some of these are trace elements (Cu, Ni, Zn) required for growth it is rare for these to be limiting and most concerns would be for their excessive application to soils.

The metal addition to soils from the different waste input types were compared by estimating the number of applications required to exceed the soil limits and the application rates with sludge regulation annual application (10 year averages). These values were then compared with the benchmark sewage sludge. Results are summarised in Table 4.2 below and presented in full in Tables A10.3 in the appendix to this report.

For most wastes the number of applications required to exceed the soil limits were similar or greater than sewage sludge indicating there was little enhanced concern. The metal which required the least number of applications to reach its soil limit was Zn where 90 applications were required for sewage sludge before reaching the limits. For some wastes the number of applications was lower. Most notable amongst these was MBT residues (code 54) which may be considered for recycling to soil. Other low application numbers to exceed Zn soil limits were from wood, de-inking sludges from paper recycling and dredgings. These would not be considered as materials applied for N fertiliser value however, although dredgings (code 73) are typically applied to soils at high loadings, and are a very variable material depending on their source.

Mean annual application rates were close to or exceeded the statutory limits for metal applications for many of the wastes indicating that there may be more concern over annual metal loading rates rather than build up of excessive soil metal concentrations.

Table 4-2 Summary of waste metal application rates that are close to or exceed sewage sludge application limits (wastes applied to PAN of 100 kg/ha)

Metal Waste loadings close to annual limit

Waste loadings exceeding annual limit

Cd Sludges from biological treatment of industrial waste MBT residues

Beverage production Construction and demolition (soil)

Cr Wood MBT residues

Construction and demolition (soil)

Cu Wood

MBT residues De-inking sludges from paper recycling Dredgings Construction and demolition (soil)

Pb Wood MBT residues Dredgings


50

Metal Waste loadings close to annual limit

Waste loadings exceeding annual limit

Construction and demolition (soil)

Hg Sludges from biological treatment of industrial waste Construction and demolition (soil)

Ni Sludges from biological treatment of industrial waste

MBT residues Dredgings Construction and demolition (soil)

Zn Wood MBT residues De-inking sludges from paper recycling Dredgings Construction

A similar analysis for the waste treatments (Table A10.4 in the Annex) indicated that again Zn was the most sensitive metal with the least number of applications required to exceed soil limits. The composts (code CL, CM and C) all showed fewer applications compared with sewage sludge with only 30 – 40 applications needed. This is a consequence of the higher application rates required to achieve the same PAN as sewage sludge, even though Zn contents of the composts (164 – 221 mg/kg DM) were lower than that of sewage sludge (675 mg/kg DM). Applications of compost where 30 applications may mean that soil metal limits would be exceeded may not be seen as sustainable.

The treatment annual application rates were all less than the statutory limits although some were close to the limits, especially the compost. Given that the analysis is based on mean values this suggests some composts may exceed the limits if applied to PAN.


51

A5. ASSESSMENT AGRONOMIC VALUE OF WASTES APPLIED AS SOIL CONDITIONERS

This section considers the loadings and agronomic benefits of applying the organic wastes to increase the soil organic carbon content by 10,000 mg C/kg soil. This assessment assumes that the addition of the wastes would improve soil quality, soil structure and other soil properties such as water holding capacity in equal measure. It would require more extensive evaluation to consider whether there was a differentiation in soil quality improvement. The loadings of nutrients have been assessed and whether these would constitute agronomic benefit or environmental risk.

A5.1 Total dry matter loadings for different wastes

Figure 5.1 shows the waste loadings as t DM/ha required to achieve the same stable increase in SOC content. This takes into account the estimated biodegradability of the wastes and indicates that for most wastes loadings would be similar to sewage sludge at about 200 kg DM/ha. These values are much higher than for application rates to achieve PAN (only 5 t DM/ha in the case of sewage sludge). A few wastes required very high applications although these are not wastes with high organic matter contents and many are applied to soil as a soil substitute (dredgings, construction and demolition soil, sludge from treatment of drinking water).

Figure 5.1 Loadings of waste types as soil conditioner

The data for the loadings by treatment (Figure 5.2) indicates that the composted materials would require similar loadings as sewage sludge. This is because although the organic matter in composts is considered stable the LOI content of composts is


52

typically much lower than for sewage sludge. Higher applications were apparent for liquid storage, not specified and AD treated wastes (R).

Figure 5.2 Loadings of waste treatments as soil conditioner

A5.2 Total N applications when applying wastes to increase SOC

The increased waste DM applications significantly increase the total N loading to soil. For the waste types many TN loadings would be at least 5,000 kg/ha (Figure 5.3) with some exceeding 10,000 kg N/ha, e.g. dairy production waste, foodwaste, MSW with manure and sludges from treatment of drinking water. Whilst many wastes would never be applied to soil to obtain an increase in SOC it would seem that TN loadings would be high. Even if a loading of 5,000 kg N/ha was considered it would need to be applied over a twenty year period in order not to exceed NVZ limits for grassland. This suggests that high application rates of organic wastes may be severely restricted on the basis of TN loadings.

The loadings of TN for the waste treatments (Figure 5.4) show a similar trend with high N loadings for all treatments although the compost loadings are about half that of the benchmark sewage sludge.


53

Figure 5.3 Total N loadings for waste types applied as soil conditioner


54

Figure 5.4 Total N loadings for waste treatments applied as soil conditioner

A5.3 Loadings of PAN for applications to increase SOC

The higher loadings of total N may not be an environmental issue if the N was stabilised and did not readily mobilise and either leach from the soil as nitrate or be available for transformation to N2O. However, total nitrogen loadings are limited by nitrogen vunerable zones (NVZ) legislation to 250 kg/ha for total nitrogen per year.

A comparison of the potentially available nitrogen (PAN) loadings has been made to provide an approximation of the risks of excessive available N for loadings to increase SOC (Figures 5.5 and 5.6. This indicates that there may be a considerable variation in the PAN loadings and therefore a corresponding wide range in risk of emissions for different wastes. For composted wastes the estimated PAN loadings are low (~500 kg N/ha) in comparison to sewage sludge (~4,000 kg N/ha). As the composted PAN is based on an assumed stabilised waste this may indicate that high loads of composted waste do not pose a significant risk of N emissions. This is an area that may require further evaluation.

Figure 5.5 Loadings of PAN for waste types applied as soil conditioner


55

Figure 5.6 Loadings of PAN for waste treatments applied as soil conditioner

A5.4 Phosphorus loadings for applications to increase SOC

Increasing the loadings to adjust the SOC content would also increase P loadings. Figures 5.7 and 5.8 indicate that these would greatly exceed plant P requirements although this assessment is based on a high P availability of 60% for all the wastes. Comparing the waste types with sewage sludge indicates that there would be a wide variability in the available P loadings and that many wastes may be applied at high application rates with a low risk of applying a significant excess of P. The green waste materials would seem to provide least risk.

The estimated P loading from composts are also much less than sewage sludge and although they would still exceed plant crop requirements the loadings may not be a significant risk especially if the percentage availability of the P is less than the assumed value used in this assessment.


56

Figure 5.7 Available P loadings from waste types applied as soil conditioner

Figure 5.8 Available P loading from waste treatments applied as soil conditioner

A5.5 Loadings of other major nutrients K, Mg and S

Applications of organic wastes as soil conditioner to increase the SOC would significantly increase the associated loadings of K, Mg and S (Figures 5.9 to 5.14).


57

The analysis indicates that this would be similar for most wastes although the amounts would greatly exceed plant growth requirements by more than 3 to 8 times even if applied over a 10 year period. It is not expected that this would be problematical for K and Mg. Of potential concern may be the high loadings of S which, if in the form of reduced S, may enhance odour and soil acidification risks. Alternatively if combined with a readily biodegradable waste there may be significant sulphate reduction causing odours and a potential adverse impact on soil quality from the toxic sulphides produced.

Composts have similar K and Mg loadings compared with sewage sludge. The impacts of excessive K, Mg and S loadings may need to be considered in more detail. For some wastes the loadings of K, Mg and S may be particularly high and these are noted in Table 5.1. The high S loadings are from some wastes that may have a high biodegradability (e.g. dairy production waste) and therefore may pose a risk from the combination of high BOD and S.

Figure 5.9 Available K for waste types applied as soil conditioner


58

Figure 5.10 Available K for waste treatments applied as soil conditioner

Figure 5.11 Total Mg loadings for waste types applied as soil conditioner


59

Figure 5.12 Total Mg loadings for waste treatments applied as soil conditioner

Figure 5.13 Total S loadings for waste types applied as soil conditioner


60

Figure 5.14 Total S loadings for waste treatments applied as soil conditioner

Table 5-1 Summary of wastes and waste treatments with high nutrient loadings when applied as soil conditioner

Nutrient Wastes with high loadings Treatment with high loadings

K Vegetable washings Vegetable production waste Pig slurry Cattle slurry Dairy production waste Biodegradable kitchen and catering waste Construction and demolition (soil)

H R U

Mg Pig slurry Dairy production waste Sugar processing Construction and demolition (soil) Fish farm waste BMW Dredgings

R

S Dairy production waste Vegetable washings Sugar processing Dredgings

N R


61

Nutrient Wastes with high loadings Treatment with high loadings

Unspecified

A5.6 Metal loadings

Tables A10.5 and A10.6 in the Annex show the number of applications required to exceed soil metal limit limits and the metal loading rate for waste applications to increase the SOC content. The larger loadings required to increase the SOC would reduce the number of applications that may be possible before exceeding the soil metal limits. This appears to be most sensitive for Zn where for many wastes only 1-4 applications would be possible before soil guideline values were breached. This would not be a sustainable practice even if the applications were spread over a longer period.

A similar picture is presented for Zn with the different waste treatments and although composts may apply less Zn than sewage sludge there may be similarly few applications (between 5 -10) before soil Zn limits are exceeded. The loadings of metals for many of the other treatments may also be high and exceed annual application rates and severely limit the number of applications required to exceed soil metal limits.

These results indicate that repeated large applications of most organic wastes to increase SOC contents may not be a sustainable practice as it may significantly increase soil metal contents (particularly Zn). A significant increase in soil metal concentrations would limit the future potential of the soil to receive more organic wastes. However, this assessment is based on total metal additions and the adverse effects of metal accumulation in soil may be related to the bioavailability and mobility of the metals which may lead to adverse biological affects but also the reduction in bioavailable metals (from crop uptake and leaching). These factors may need to be considered in more detail if total metal applications are significant.


62

A6. GASEOUS EMISSIONS TO AIR AND CARBON SEQUESTRATION

A6.1 Introduction

The application of the wastes for fertiliser and as soil conditioner would contribute to the risk of gaseous emissions from soil, including greenhouse gases, but would also allow for carbon sequestration in soil as a means of mitigating global warming. These factors are considered in this section

A6.2 Emissions of ammonia to air

A6.2.1 Ammonia emissions from applications as fertiliser to PAN

The results of the estimated NH3 emissions from applying the waste types are shown in Figure 6.1.

The available information on NH3 contents in the organic wastes is limited and therefore the assessment of NH3 emissions is considered as an initial assessment that could be updated when more reliable waste information is available. However, for most wastes where data is available NH3 emissions to air are lower or similar to that of sewage sludge. Where emissions may exceed that of sewage sludge the levels may only be slightly higher by 2-3 times. The potentially high NH3 emitting wastes are manures, sludge from treatment of industrial effluents, ABPR waste and de-inking sludges from paper recycling.

Insufficient data was available for a comparison of emissions to air by waste treatment although composts typically contain comparatively little free NH3 compared to sewage sludge and manures and lower emissions would be expected to be lower than sewage sludge. By assuming all the mineral N in compost (code C) is NH3 then the emissions are estimated as about 5 times lower for composts compared with sewage sludge loaded to the same PAN.

The NH3 emissions from application of aerobic digestion solid and liquid digestates are not known but would be expected to be the same order as sewage sludge and manures due to their high NH3 content.

A6.2.2 Ammonia emissions from applications as soil conditioner

When the wastes are added at higher application rates as soil conditioners to increase the SOC content then there are significant increases in potential NH3 emissions (Figure 6.2). For example the estimated NH3 emissions from sewage sludge increase from about 20 to 880 kg N/ha. Some of the wastes may give higher NH3 emissions than sewage sludge (e.g. pig slurry, food waste, sludges from washing and cleaning for food production, MSW and manure, BMW, dairy production waste). Green waste however is seen as presenting a low risk of NH3 emission.

As with the loadings to PAN there is no data for the treatments when applied as soil conditioner. However if the same assumption is made regarding the NH3 content of composts then the emissions from composts would only be around 28 kg N/ha compared with sewage sludge at 880 kg N/ha.


63

Figure 6.1 Ammonia emissions from application of waste types to PAN of 100 kg N/ha

Figure 6.2 Ammonia emissions from waste types applied as soil conditioner

A6.2.3 Emissions of nitrous oxide

Emissions of N2O from applications of organic wastes will be very difficult to predict as they are dependent on a multitude of factors such as the waste characteristics


64

and its decomposition, the microbial processes of nitrification and denitrification in the soil and the factors such as moisture, oxygen, and temperature that affect these process (Section A2.4.2). For a fist pass comparison 1% of the applied N is assumed to be released as N2O which in practice takes a long term view of the N applied. Actual N2O emissions would be time dependent and may be different for different wastes in the short term as the more readily available (PAN) nitrogen is metabolised within the soil. The assessment would present a similar pattern to the total N applications and it is considered that the assessment may be refined if more specific short term data was available.

A6.2.4 Emissions of nitrous oxide for application to PAN

The results (Figures 6.3 and 6.4) indicate that N2O emissions may be similar relative to sewage sludge with only a few wastes that may give greater emissions (e.g. sludges from biological treatment of industrial effluents, ABPR wastes, MBT residues, and de-inking sludges from paper recycling).

The analysis of N2O emissions by waste treatment indicates that composted materials may give about 5 times more emissions relative to sewage sludge. These high estimations result from the high loadings of compost to give the same PAN as sewage sludge which results in a corresponding high total N application rate. As this N is relatively stable it illustrates the key question of whether this stabilised material N poses the same risk in overall N2O emissions as the more biologically available N of materials such as sewage sludge and digestates. Further work is recommended to resolve this issue.

Figure 6.3 Nitrous oxide emissions from waste types applied to PAN of 100 kg N/ha


65

0.00

2.00

4.00

6.00

8.00

10.00

12.00


Nit

rou

s o

xid

e e

mis

sio

n (kg

N/h

a)

Waste treatment

Figure 6.4 Nitrous oxide emission from waste treatment applied to PAN of 100 kg N/ha

A6.2.5 Nitrous oxide emissions from application as soil conditioner

The potential N2O emissions when applied as soil conditioner (Figure 6.5 and 6.6) would be significantly higher than when applied as fertiliser. For example the estimated emission from sewage sludge was 2 kg N/ha when applied as N fertiliser and 87 kg N/ha if applied as soil conditioner. Potential N2O emissions from compost applications as soil conditioners would be lower than for sewage sludge. This is because of the lower total N loaded to the soil (Section A5.2).


66

Figure 6.5 Nitrous oxide emissions from waste types applied as soil conditioner

Figure 6.6 Nitrous oxide emissions from waste treatments applied as soil conditioner


67

A6.3 Carbon dioxide emissions and carbon sequestration

Carbon dioxide is a greenhouse gas that would be produced from the decomposition of the organic waste in soil and emitted to atmosphere. In this assessment it is assumed the extent of decomposition of the organic waste in the soil and therefore the emission of CO2 is based on this assumed decomposition.

A6.3.1 Carbon dioxide emissions from applications to PAN

Figures 6.7 and 6.8 show the CO2 emissions from applications of waste and waste treatments to PAN. The results indicate that some waste types would contribute higher CO2 emissions (e.g. sludges from biological treatment of industrial waste, ABPR waste and de-inking sludges from paper recycling). The CO2 emissions from the composted wastes are similar to that predicted for sewage sludge as despite the waste carbon being more stable, the applied organic C is much greater.

Figure 6.7 CO2 emissions from waste types applied to PAN of 100 kg N/ha


68

Figure 6.8 CO2 emissions from waste treatments applied to PAN of 100 kg N/ha

A6.3.2 Carbon dioxide emissions when applied as soil conditioner

The corresponding CO2 emissions where wastes are applied as soil conditioner are shown in Figures 6.9 and 6.10. The assessment results are dependent on the assumed biodegradability of the waste in the soil. Hence the CO2 emitted by manures is less than half that of sewage sludge even though it is assumed that the manures are only slightly less biodegradable (40%) than sewage sludge (60%). This is because as the OC is less biodegradable in manures correspondingly less is required to provide the target increase in stable SOC. Therefore the amount that is decomposed and emitted as CO2 is significantly lower. The CO2 emissions from composts where the organic C is stabilised are much less than sewage sludge.


69

Figure 6.9 CO2 emissions of waste types applied as soil conditioner

Figure 6.10 CO2 emissions from waste treatments applied as soil conditioner


70

A6.3.3 Carbon sequestration

The carbon sequested is the converse of the estimated CO2 emission. For the applications as soil improver these were defined as 30 t C/ha for all wastes as the normalisation method. The carbon sequested for applications to PAN are shown in Figures 6.11 and 6.12. These results indicate the high C sequestration expected from composted organic wastes. In this context it is also important to consider the long term impacts of organic matter applications to soil on carbon sequestration. For example improvements in soil structure may increase the amount of C sequestrated in soil from the growth of crops.

This has not been considered here but potentially models for this may be included within the approach used here to provide a more reliable prediction.

Figure 6.11 Estimated C sequested in soil from application of waste types to PAN of 100 kg n/HA


71

Figure 6.12 Estimated C sequested in soil from application of waste treatments to PAN of 100 kg N/ha

A6.3.4 Methane Emissions

The assessment approach is not readily applicable to the estimation of CH4 emissions from soils which are generally assumed to be a major sink of atmospheric CH4 by the oxidation of methane oxidising bacteria in the soil. The addition of high levels of biodegradable organic waste may increase the production of CH4 in soils and limit the action of CH4 oxidising bacteria through the creation of anoxic conditions. This may conceivably turn soils into a CH4 source rather than a sink. The risk of this may match the CO2 emissions in the worst case scenario but the initial biological demand of the decomposing waste will have a greater affect. Therefore to fully predict this aspect data on the both extent and rate of organic waste decomposition in soil may be required. A further issue of concern would be whether any or a combination of parameters added to soil in the waste was inhibitory to the methane oxidising bacteria as this would adversely impact soils as a sink of atmospheric methane.


72

A7. ORGANIC POLLUTANTS

There is a whole multitude of potential organic pollutants of concern and data for their concentrations in most organic wastes is sparse. Often data is based on chemical classes and at best for a select few classes of compounds although these may be of high concern. Typically such classes may be:-

TPH – total petroleum hydrocarbons;

PCB – poly chlorinated biphenyls;

PAH – poly aromatic hydrocarbons;

PCDD/F- Polychlorinated dibenzo-p-dioxins and furans;

LAS – linear Alkylbenzene Sulfonates;

DEHP – di(2-ethyylhexyl)phthalate;

Pesticides;

AOX – chlorinated organic compounds.

The fate of the pollutant once in the soil environment is complex as is its potential risk of causing harm. For example solubility in water, adsorption to other soil materials, microbial decomposition, accumulate in crops, and leaching into surface and groundwaters are all important factors to consider.

Being a potential C and energy source, organic pollutants may be decomposed in soils by soil micro-organisms. However, some organic pollutants are resistant to microbial decomposition, and some organic pollutants although biodegradable may be below threshold concentrations for the microbes to recognize the pollutant as a valuable carbon source to decompose.

For many organic pollutants the toxic or other biological effect such as endocrine disruption may occur at very low concentrations. These low concentrations may be below microbial threshold concentrations for decomposition. In this case the widespread application of wastes to soils may be a risk that requires further consideration especially as a source of diffuse pollution which may only become an issue many years after application, e.g. if the diffuse pollution enters groundwaters used as potable water sources.

The methodology used for assessing the loadings of other parameters may also be applied to organic pollutants and compared with any limits for soil organic pollutants. Where such limits have been derived through a risk assessment approach as in the EA CLEA (EA 2009b) model then the limits may be used with some justification. However, few soil limit values have been derived to date.

In a worst case scenario the number of applications of any waste may be estimated that results in the soil organic pollutant level being exceeded. This would be a similar approach to the metals assessment and would also assume there were no losses from microbial decomposition.


73

A7.1 PAH and PCB assessment for waste applied as soil conditioners.

Data from various sources on the PAH and PCB content of several wastes (Table 7.1) has been collated. These have then been used to estimate the number of applications to exceed example soil concentrations of 3 and 0.05 mg/kg from an initial typical soil concentration of 0.05 and 0.01 mg/kg for PAH and PCB respectively. These soil concentrations and limits were selected for illustration purposes and actual appropriate values would need to be derived for a robust assessment methodology.

The results indicate that applications for some wastes may significantly increase soil organic pollutant concentrations after only a few applications. This approach may be used as more data on wastes becomes available and the accumulation may be modified to account for biodegradation to obtain a more reliable assessment. Further work is therefore recommended to improve the understanding of organic pollutants in soil from organic waste applications to support this development.

Table 7-1 Example analysis of impact of selected wastes applied as soil conditioner on soil organic pollutant levels

Waste PAH mg/kg DM

Applications to

soil limit PCB mg/kg DM

Applications to

soil limit

Sewage sludge 0.5 79 0.001 53

Sewage sludge 10 3.9 0.23 0.1

Pig slurry 0.05 1250 0.0035 242

Compost Low 0.8 42 0.03 15

Compost medium 2.4 14 0.75 0.6

Compost high 9.7 3 1.68 0.15


74

A8. GROUND AND SURFACE WATER RISK ASSESSMENT

A8.1 Introduction

The application of organic materials to land has the potential to impact on both surface and groundwaters through leaching and direct run-off. An assessment of risk to surface and groundwater from the waste types was carried out using Consim® for the generic waste groups previously identified.

A8.2 Methodology

The availability of leachability data is extremely limited and the assessment of risk to ground and surface water was therefore undertaken based on chemical composition. The Consim® model developed by Golders Associates for the Environment Agency was used to assess the risk to of applying organic materials to agricultural land.

The software utilises a tiered assessment using successive levels. Level one assumes no dilution or attenuation of the contamination and is thus the most conservative of the three potential levels of assessment. Further levels require additional information on the application site and are less conservative.

Consim Level two analyses the behaviour and travel time in the unsaturated zone. Given the large variability of hydrogeological conditions across England and Wales and of the unsaturated zone thickness, it was not possible to develop a generic methodology for a Level two groundwater risk assessment. A Level one risk assessment using Consim® has therefore been undertaken for the parameters identified in Table 8-1.

Table 8-1 Statutory benchmarks used for risk assessment

Surface water benchmark Groundwater benchmark

Chemical Value (mg/l) Source Value (mg/l) Source

Nitrogen 1# Freshwater Fish Directive

0.5 WS (WQ) Reg. 2000 (SI 3184)

Phosphorus 0.1 EA No standard

Extractable Phosphorus

0.04$* EC report No standard

Potassium No standard No standard

Zinc 0.008* EQS EW 2007 3 WHO 2006

Copper 0.0005* EQS EW 2007 2 DWI (E&W2007)

Nickel 0.008* EQS EW 2007 0.02 DWI (E&W2007)

Lead 0.004* EQS EW 2007 0.025 DWI (E&W2007)

Cadmium 0.005 EQS EW 2007 0.005 DWI (E&W2007)

Chromium 0.002* EQS EW 2007 0.05 DWI (E&W2007)


75

Surface water benchmark Groundwater benchmark

Chemical Value (mg/l) Source Value (mg/l) Source

Mercury 0.001 EQS EW 2007 0.001 DWI (E&W2007)

Note: All values are based on an annual average. * indicate that the value depends on the hardness of the river (concentration in CaCO3). The lowest concentration within hardness bands has been selected. # indicates Total ammonium and is imperative standard in Freshwater Fish Directive.

$ has been selected from UK report to the EC in order to achieve “good/moderate status” of the water body by the

WFD.

As there was no leaching test information available for many of the waste materials, theoretical leachate concentrations have been calculated by Consim from the soil concentration data and chemical testing data for each waste type. The calculation incorporates theoretical solid/liquid/gas partition effects and a mass balance:

RcHqaqwKd

CsCl

/)*(

Where:

Cl = leachate concentration (mg/l)

Cs = soil concentration (mg/kg)

Kd = partition coefficient for contaminated soil (ml/g)1

qw = water filled porosity of contaminated soil (fraction)

qa = air filled porosity of contaminated soil (fraction)

H = Henry’s Law constant (dimensionless)

Rc = contaminated soil dry bulk density (g/cm3)

ConSim compares the calculated leachate concentration (Cl) with the maximum solubility of the contaminant. Concentrations in excess of the maximum solubility will be limited to the maximum solubility. Definition of the source (summarised in

Table 8-2) includes dry bulk density, air filled porosity, water filled porosity, thickness and dimension of the source (length and width) from loading rate information and these parameters are estimated from description and analysis of the different waste types. An area of 1 hectare (100m by 100m square) has been selected to reflect the risk to the definition of application rate (tonnes of materials disposed to land per hectare). As discussed in Section A3, the waste disposal to land will transform the physico-chemical characteristics of the soil and will affect the leachable potential of the soil. However, it is important in the groundwater and surface water risk assessment to identify the quality of the waste applied to land itself. The waste disposed to land will have chemical characteristics of the waste and will be input in the Consim model as being typical UK soil physical characteristics as identified below.

1 Where actual values of Kd are are not available, these have been calculated as partition to organic

carbon (Koc) multiplied by the soil fraction of organic carbon (foc), e.g. for N, K and P.


76

Table 8-2 Soil characteristics used for risk assessment

Parameter Value Unit

Dry bulk density 1.5 g/cm3

Air filled porosity 40 %

Water filled porosity 20 %

Incorporation depth 15 cm

Length 100 m

Width 100 m

The wide range of application rates that have been identified previously cannot be applied for each waste type because application rates will vary with objective for the specific soil (achieve PAN of 100 kg N/ha or increase in organic content of the soil). The incorporation depth will also vary according to regional practice and soil type (tillage depth can vary from 15 cm to 50 cm). The methodology taken for the surface and groundwater risk assessment considers that the application of contaminated material should be prevented, and therefore 100% of waste is used in this risk assessment without any dilution in the soil. The leachate chemical composition produced by the Consim model is compared against relevant benchmark without considering any attenuation factor. It then implies that a direct contamination of either groundwater or surface water is achieved through leaching of the waste applied.

Because Consim® is a probabilistic model, the uncertainty on the concentration of chemicals can be represented by a distribution function representing the variability of the concentration for that chemical. From this distribution function, a minimum value, a most likely value and a maximum value can be derived.

A Level one assessment was carried out of risk to surface and groundwater of organic materials spread to land. Thirteen chemical parameters have been identified for the risk assessment for surface and groundwater as detailed in Table 8-1.

The following parameters have been assessed:

nitrogen (N) and available nitrogen;

phosphorus (P) and available phosphorus;

potassium (K) and available potassium;

and metals (zinc, copper, nickel, lead, cadmium, chromium, mercury).

Kd has been calculated using Koc and organic content in agricultural soil (Comparison of original and re-sampled National Soil Inventory data, MAFF project SP0506 20002)

2http://randd.defra.gov.uk/Default.aspx?Menu=Menu&Module=More&Location=None&ProjectID=9320&FromSearch=

Y&Publisher=1&SearchText=sp05&SortString=ProjectCode&SortOrder=Asc&Paging=10#Description

http://randd.defra.gov.uk/Default.aspx?Menu=Menu&Module=More&Location=None&ProjectID=9320&FromSearch=Y&Publisher=1&SearchText=sp05&SortString=ProjectCode&SortOrder=Asc&Paging=10#Description

http://randd.defra.gov.uk/Default.aspx?Menu=Menu&Module=More&Location=None&ProjectID=9320&FromSearch=Y&Publisher=1&SearchText=sp05&SortString=ProjectCode&SortOrder=Asc&Paging=10#Description


77

for N, P and K. A triangular distribution function for the organic content of soil (foc) has been estimated from that data (minimum = 0.43%; median = 2.3% and maximum is 46.81%) and has been multiplied by Koc to obtain Kd (Koc*foc=Kd). For N and P the Koc value is 14 l.kg-1, and 35 l.kg-1 for potassium. For metals, Kd values have been selected from literature (EA Science Report SC050021 / Mercury soil guidline value (SGV), EA Science report: SC050021 / Nickel SGV, USEPA – Understanding variation in partition coefficient, Kd values volume 1 and 2). As Kd is largely dependent on pH, it has been assumed that range of pH values would vary between 6 and 8 which would cover all but the most acidic soil types in England and Wales.

A8.3 Concentration input

The input concentrations for all 13 parameters assessed for this exercise are described in Table 8.3, and details on the distribution function are given for each parameter. A summary table of the 95%ile value for each waste group is given in Table 8-3. The detailed cumulative functions have been discussed in this section for each waste type and for each parameter (Figure 8.1 to Figure 8.13) presented in Section A8.3.1 to Section A8.3.10.

In line with current European and UK practice for the derivation of parameter values for human health, and environmental risk assessment as detailed in Carlton’s report (2007), 95th percentiles have been used ofr each dataset3. The overall waste parameters concentration indicates relative concentration between waste type for each parameters. Table 8-3 shows that relative high concentrations of nutrient (N, P and K) are mostly contained in manures, sewage sludge and food waste. Manures and sewage sludge also contains concentrations of Zn and Cu respectively ten times higher than greenwaste, biowaste and the other waste group. Priority hazardous substances (Cd and Hg) are extremely high in foodwaste and sewage sludge when compared with other groups and have been statistically well assessed (the number of value is respectively 426 and 1305). However, the range of values for foodwaste varies significantly within the group and a more detailed analysis should identify the most significant contributors to such high levels.

3 Carlon, C. (Ed.) (2007). Derivation methods of soil screening values in Europe. A review and

evaluation of national procedures towards harmonization. European Commission, Joint Research Centre, Ispra, EUR 22805-EN, 306 pp


78

Table 8-3 Summary table of 95%ile concentrations (mg/kg DM) used in the groundwater and surface water risk assessment

Concentration used in risk assessment (95%ile value)

Waste type Total N Extract N Total P Extract P Total K Extract K Zn Cu Ni Pb Cd Cr Hg

Others 12000 25 1680 503 2859 1090 238 143 33 54 1.8 33 0.5

OFMSW 78685 No data 6860 1294 14750 8474 1160 582 236 748 4.1 510 1.4

Biowaste (mixed organic wastes) 64115 2440 10750 1352 39585 6492 449 165 105 136 1.0 42 0.25

Foodwaste 110638 9585 20819 11482 33897 35541 955 585 280 117 22 500 26

Greenwaste 23965 760 3509 184 13492 6758 323 122 27 228 1.2 50 0.5

Sewage waste 70145 16288 43790 14550 11992 322 1373 903 99 387 5.3 239 5.5

Manures 171829 2760 36517 3453 69620 27565 2457 939 26 19 1.1 37 0.07

All 77000 983 36000 4891 28333 9075 1090 766 97 323 5.8 198 6.1

Number of analysis 3022 348 2971 491 1708 407 3163 3168 3022 3009 2829 2998 2813


79

The following series of plots show the data distribution for each parameter. The x axis gives the concentration of the determinand, and the y axis gives the cumulative frequency as a proportion of the total.

A8.3.1 Nitrogen

Total Nitrogen

Cumulative frequency of total nitrogen for different waste types

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1

1.E+00 1.E+01 1.E+02 1.E+03 1.E+04 1.E+05 1.E+06

Concentration (mg/kg DM)

Cu

mu

lati

ve

fre

qu

en

cy

(fr

acti

on

)

Others

OFMSW

Biowaste (mixed organic wastes)

Foodwaste

Greenwaste

Sewage waste

Manures

All types

Figure 8.1 Total nitrogen concentration in different waste types


80

Extractable nitrogen

Cumulative frequency of extractable nitrogen for different waste types

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1

1.E-02 1.E-01 1.E+00 1.E+01 1.E+02 1.E+03 1.E+04 1.E+05


Cu

mu

lati

ve

fre

qu

en

cy

(fr

acti

on

)

Others

OFMSW


Foodwaste

Greenwaste

Sewage waste

Manures

All types

Figure 8.2 Extractable nitrogen concentration in different waste types

A8.3.2 Phosphorus

Total phosphorus

Cumulative frequency of total phosphorus for different waste types

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1

1.E+00 1.E+01 1.E+02 1.E+03 1.E+04 1.E+05 1.E+06


Cu

mu

lati

ve

fre

qu

en

cy

(fr

acti

on

)

Others

OFMSW


Foodwaste

Greenwaste

Sewage waste

Manures

All types

Figure 8.3 Total phosphorus concentration in different waste types


81

Available phosphorus

Cumulative frequency of extractable phosphorus for different waste types

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1

1.E-01 1.E+00 1.E+01 1.E+02 1.E+03 1.E+04 1.E+05


Cu

mu

lati

ve

fre

qu

en

cy

(fr

acti

on

)

Others

OFMSW


Foodwaste

Greenwaste

Sewage waste

Manures

All types

Figure 8.4 Available phosphorus concentration in different waste types

A8.3.3 Potassium

Total potassium

Cumulative frequency of total potassium for different waste types

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1

1.E+00 1.E+01 1.E+02 1.E+03 1.E+04 1.E+05 1.E+06


Cu

mu

lati

ve f

req

uen

cy (

fracti

on

)

Others

OFMSW


Foodwaste

Greenwaste

Sewage waste

Manures

All types

Figure 8.5 Total potassium concentration in different waste types


82

Available potassium

Cumulative frequency of extractable potassium for different waste types

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1

1.E-01 1.E+00 1.E+01 1.E+02 1.E+03 1.E+04 1.E+05


Cu

mu

lati

ve

fre

qu

en

cy

(fr

acti

on

)

Others

OFMSW


Foodwaste

Greenwaste

Sewage waste

Manures

All types

Figure 8.6 Available potassium concentration in different waste types

A8.3.4 Zinc

Cumulative frequency of zinc for different waste types

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1

1.E-02 1.E-01 1.E+00 1.E+01 1.E+02 1.E+03 1.E+04


Cu

mu

lati

ve

fre

qu

en

cy

(fr

acti

on

)

Others

OFMSW


Foodwaste

Greenwaste

Sewage waste

Manures

All types

Figure 8.7 Zinc concentration in different waste types


83

A8.3.5 Copper

Cumulative frequency of copper for different waste types

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1

1.E-02 1.E-01 1.E+00 1.E+01 1.E+02 1.E+03 1.E+04


Cu

mu

lati

ve

fre

qu

en

cy

(fr

acti

on

)

Others

OFMSW


Foodwaste

Greenwaste

Sewage waste

Manures

All types

Figure 8.8 Copper concentration in different waste types

A8.3.6 Nickel

Cumulative frequency of nickel for different waste types

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1

1.E-02 1.E-01 1.E+00 1.E+01 1.E+02 1.E+03 1.E+04


Cu

mu

lati

ve

fre

qu

en

cy

(fr

acti

on

)

Others

OFMSW


Foodwaste

Greenwaste

Sewage waste

Manures

All types

Figure 8.9 Nickel concentration in different waste types


84

A8.3.7 Lead

Cumulative frequency of lead for different waste types

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1

1.E-02 1.E-01 1.E+00 1.E+01 1.E+02 1.E+03 1.E+04


Cu

mu

lati

ve

fre

qu

en

cy

(fr

acti

on

)

Others

OFMSW


Foodwaste

Greenwaste

Sewage waste

Manures

All types

Figure 8.10 Lead concentration in different waste types

A8.3.8 Cadmium

Cumulative frequency of cadmium for different waste types

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1

1.E-02 1.E-01 1.E+00 1.E+01 1.E+02 1.E+03 1.E+04


Cu

mu

lati

ve

fre

qu

en

cy

(fr

acti

on

)

Others

OFMSW


Foodwaste

Greenwaste

Sewage waste

Manures

All types

Figure 8.11 Cadmium concentration in different waste types


85

A8.3.9 Chromium

Cumulative frequency of chromium for different waste types

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1

1.E-02 1.E-01 1.E+00 1.E+01 1.E+02 1.E+03 1.E+04


Cu

mu

lati

ve

fre

qu

en

cy

(fr

acti

on

)

Others

OFMSW


Foodwaste

Greenwaste

Sewage waste

Manures

All types

Figure 8.12 Chromium concentration in different waste types

A8.3.10 Mercury

Cumulative frequency of mercury for different waste types

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1

1.E-02 1.E-01 1.E+00 1.E+01 1.E+02 1.E+03 1.E+04


Cu

mu

lati

ve

fre

qu

en

cy

(fr

acti

on

)

Others

OFMSW


Foodwaste

Greenwaste

Sewage waste

Manures

All types

Figure 8.13 Mercury concentration in different waste types


86

A8.3.11 Results of the risk assessment

Table 8-4 summarises the resulting 95%ile leachate concentration given by the Consim model for each waste category.

The result expressed in 95%ile concentration represents the leachate concentration that is not exceeded for 95% of the calculated values. 5000 iterations have been used during the Monte Carlo analysis and therefore a 95%ile concentration represents the maximum value achieved for 4750 calculated leachate concentrations. 250 calculated concentrations exceed this concentration.


87

Table 8-4 Summary table of 95%ile leachate concentrations (mg/l) calculated by Consim for the groundwater and surface water risk assessment with benchmark used.

Resulting leachate concentration calculated by Consim (95%ile value)

Waste type Total N Extract N Total P Extract P Total K Extract K Zn Cu Ni Pb Cd Cr Hg

Others 339 0.7 47.29 13.84 31.72 11.48 0.47 0.29 0.066 0.108 0.00773 20.20 0.000094

OFMSW 2278 No data 184.68 37.75 167.17 88.72 2.32 1.16 0.472 1.495 0.01707 323.83 0.000275


1812 66.6 310.86 39.58 451.72 457.57 0.90 0.33 0.209 0.333 0.00456 26.20 0.000049

Foodwaste 3131 262.0 593.88 339.65 399.43 415.64 1.91 1.17 0.560 0.233 0.09776 313.81 0.005031

Greenwaste 704 22.4 99.01 5.18 153.14 77.00 0.65 0.24 0.054 0.456 0.00610 31.81 0.000092

Sewage waste 1975 443.2 1200.07 417.67 133.55 3.63 2.75 1.81 0.198 0.773 0.02327 149.76 0.001000

Manures 4984 75.1 1000.75 99.11 791.31 311.48 4.91 1.88 0.052 0.038 0.00498 23.36 0.000013

Benchmark* 1 1 1 0.04 N/A N/A 3 2 0.02 0.025 0.005 0.05 0.001

* Benchmarks selected are most stringent EQS (depending on water hardness for N, P and K) and groundwater standards fro metals (in mg/l).


88

A8.4 Discussion and conclusion

Maximum 95%ile leachate concentrations are presented with typical water quality benchmark (EQS and DWS) along with groundwater background concentrations in Table 8-5. A more detailed analysis for each waste type is given below.

Metals

Regardless of whether a material is considered to be a waste, the requirements of the Groundwater Regulations 1998 (and Water Framework Directive and its Daughter Directive on Groundwater) still apply. The entry of priority hazardous substances (also known as substances in List I) into groundwater must be prevented (unless certain exemptions apply) and the introduction of priority substances (known as substances in List II) must be limited so as to avoid pollution. Cadmium (and its compounds) and mercury (and its compounds) are identified as priority hazardous substances (List I). Nickel (and its compounds) and lead (and its compounds) are priority substances (List II).

Priority hazardous substances are present in the list of wastes reviewed for this study albeit at relatively low concentrations. It is considered that these substances do not present a significant risk to the environment (surface and ground waters). However, it should be noted that some material identified in this study have concentrations which are much higher than background concentrations or environmental quality standards for surface or groundwater, including sewage waste (type 20) and food waste (type 40) and may reflect episodic contamination with contaminated material. Food waste notably has concentration around 50 times more than prescribed water quality standards for these two parameters.

Modelled leachate concentrations of the priority substances Ni and Pb were 10 to 20 times higher than background concentrations for land application of biowastes (10 times), food wastes and OFMSW wastes (more than 20 times). This may pose an environmental risks if sufficient mitigation measures are not put in place.

The leachate concentrations of Zn and Cu are below respective benchmarks for all wastes, with the exception of manures for Zn. The water quality standard for Cr is above the benchmark by several orders of magnitude for all waste types identified in this study.

Mitigation measures such as effective control of the waste stream and regular chemical analysis for these elements would allow an acceptable level of contaminants, eventual blending or reduced amount of materials to be applied per hectare. However, results discussed above relate to a Tier 1 risk assessment. If sufficient data is collected to allow more detailed risk assessment using Tier 2 and Tier 3 of the model, a reduction of the overall risk to surface water and groundwater may be achieved, as these take into account parameters such as dilution and chemical reactions thus reducing effective parameter concentrations.

Nutrients: nitrogen, phosphorus and potassium

Phosphorus and nitrogen levels in England and Wales are already high, and many water bodies are failing to achieve good status. High concentrations of nutrients are found in most agronomic wastes reviewed in this study, as would be expected of material applied for agronomic benefit. Land application of organic waste with high


89

nitrogen and phosphorus content may present an obstacle for England and Wales to achieve good status for its water bodies and consider application with extreme care.

Because of the lack of benchmarks and concentrations close to background levels, potassium concentrations found in this study are not considered to present a risk to surface or groundwater for materials assessed for this study.

Table 8-5 Table comparing highest 95%ile leachate concentration with surface and groundwater threshold concentrations (EQSs and DWLs).

Chemical parameter

Highest 95%ile concentration (mg/l) for selected waste

(indicated in bracket)

Surface water# Groundwater

EQS DWL Background concentration

Total Nitrogen 171829.3 (10) 1 0.5 1.12

Extractable Nitrogen

16287.9 (20) N/A N/A 1.12

Total Phosphorus

43790 (20) 0.1 N/A 0.48

Extractable Phosphorus

14550.1 (20) 0.04 N/A 0.48

Total Potassium

69620 (10) N/A N/A 27

Extractable Potassium

35541.3 (40) N/A N/A 27

Cadmium 21.59 (40) 0.005 0.005 0.0025

Chromium 510.0 (60) 0.002 0.05 0.005

Copper 939.35 (10) 0.0005(1) (0.012) 2 0.093

Lead 747.55 (60) 0.004(1) (0.02) 0.025 0.028

Mercury 26.32 (40) 0.001 0.001 0.0002

Nickel 280.20 (40) 0.008(1) (0.04) 0.02 0.015

Zinc 2457.40 (10) 0.008(1) (0.05) 3 0.373

# Variability of surface water “natural” concentration is such that no background concentration could be easily considered unless than on a case specific basis. * 97.7%ile background concentrations are taken from “The natural (baseline) quality of groundwater in England and Wales , Research report RR/07/06 and Technical report NC/99/74/24 published by BGS and EA in 2007”. (1) EQSs selected are the most stringent EQSs for very low CaCO3 concentration (hardness <50mg/l CacO3) in rivers. The maximum concentration when considering highest hardness value (CaCO3 concentration > 250mg/l) is given in bracket for those parameters.

A8.5 Comparison of treatment method for sewage sludge

The sewage waste dataset has the most comprehensive information on treatment types and allows a comparison of different treatment type options and their relative


90

influence on the chemical composition of sewage sludge. However, care should be taken in the interpretation of the treatment method effectiveness as some sewage sludge treatment could apply to a specific type of sewage originating from various origins. The amount of industrial contribution may vary as well as the level of treatment that is achieved by each sewage treatment works.

A8.5.1 Comparison between treatment type in wet weight

Results (Figure 8.14 to Figure 8.22) are given below in two graphics per parameters comparing:

untreated (U),

mesophilic AD (D),

thermophilic AD (T),

composting (C),

mesophilic and mechanical dewatering (DM)

mechanical dewatering (M)

heat dried (H),

lime stabilisation (L)

Liquid storage (LS)


91

Sewage sludge treatment type comparison

0

5

10

15

20

25

30

35

40

45

50

10 100 1000 10000 100000

Total Nitrogen Concentration in mg/kg WW

Oc

cu

rre

nc

e (

%)

U

D

T

C

DM


0

5

10

15

20

25

30

35

10 100 1000 10000 100000


Oc

cu

rre

nc

e (

%) U

M

DM

H

L

LS

Figure 8.14 Total nitrogen concentration by treatment type


92


0

5

10

15

20

25

30

35

40

45

50

1 10 100 1000 10000 100000

Total Phosphorus Concentration in mg/kg WW

Oc

cu

rre

nc

e (

%)

U

D

T

C

DM


0

5

10

15

20

25

30

35

40

1 10 100 1000 10000 100000


Oc

cu

rre

nc

e (

%) U

M

DM

H

L

LS

Figure 8.15 Total phosphorus concentration by treatment type


93


0

10

20

30

40

50

60

70

80

90

100

0.001 0.01 0.1 1 10 100

Nickel Concentration in mg/kg WW

Oc

cu

rre

nc

e (

%)

U

D

T

C

DM


0

10

20

30

40

50

60

70

80

90

100

0.001 0.01 0.1 1 10 100

Nickel Concentration in mg/kg WW

Oc

cu

rre

nc

e (

%) U

M

DM

H

L

LS

Figure 8.16 Nickel concentration by treatment type


94


0

10

20

30

40

50

60

70

80

90

100

0.01 0.1 1 10 100 1000

Lead Concentration in mg/kg WW

Oc

cu

rre

nc

e (

%)

U

D

T

C

DM


0

10

20

30

40

50

60

70

0.01 0.1 1 10 100 1000 10000


Oc

cu

rre

nc

e (

%) U

M

DM

H

L

LS

Figure 8.17 Lead concentration by treatment type


95


0

10

20

30

40

50

60

70

0.001 0.01 0.1 1 10 100

Cadmium Concentration in mg/kg WW

Oc

cu

rre

nc

e (

%)

U

D

T

C

DM


0

10

20

30

40

50

60

70

0.001 0.01 0.1 1 10 100

Cadmium Concentration in mg/kg WW

Oc

cu

rre

nc

e (

%) U

M

DM

H

L

LS

Figure 8.18 Cadmium concentration by treatment type


96


0

20

40

60

80

100

120

0.01 0.1 1 10 100 1000

Chromium Concentration in mg/kg WW

Oc

cu

rre

nc

e (

%)

U

D

T

C

DM


0

20

40

60

80

100

120

0.01 0.1 1 10 100 1000

Chromium Concentration in mg/kg WW

Oc

cu

rre

nc

e (

%) U

M

DM

H

L

LS

Figure 8.19 Chromium concentration by treatment type


97


0

10

20

30

40

50

60

70

80

90

100

0.001 0.01 0.1 1 10

Mercury Concentration in mg/kg WW

Oc

cu

rre

nc

e (

%)

U

D

T

C

DM


0

10

20

30

40

50

60

70

80

90

100

0.001 0.01 0.1 1 10

Mercury Concentration in mg/kg WW

Oc

cu

rre

nc

e (

%) U

M

DM

H

L

LS

Figure 8.20 Mercury concentration by treatment type


98


0

5

10

15

20

25

30

35

40

45

0.1 1 10 100 1000

Zinc Concentration in mg/kg WW

Oc

cu

rre

nc

e (

%)

U

D

T

C

DM


0

5

10

15

20

25

30

35

40

45

0.1 1 10 100 1000 10000

Zinc Concentration in mg/kg WW

Oc

cu

rre

nc

e (

%) U

M

DM

H

L

LS

Figure 8.21 Zinc concentration by treatment type


99


0

5

10

15

20

25

30

35

40

45

0.1 1 10 100 1000

Copper Concentration in mg/kg WW

Oc

cu

rre

nc

e (

%)

U

D

T

C

DM


0

5

10

15

20

25

30

35

40

0.1 1 10 100 1000

Copper Concentration in mg/kg WW

Oc

cu

rre

nc

e (

%) U

M

DM

H

L

LS

Figure 8.22 Copper concentration by treatment type


100

A8.5.2 Comparison between treatment types and wet weight and dry matter

Plots for total nitrogen (Figure 8.23), total phosphorus (Figure 8.24) and lead (Figure 8.25) indicates a diminution of one order of magnitude between wet weight and dry matter.


0

5

10

15

20

25

30

35

10 100 1000 10000 100000 1000000

Total Nitrogen Concentration in mg/kg DM

Oc

cu

rre

nc

e (

%)

U

D

T

C

DM


0

5

10

15

20

25

30

35

40

45

50

10 100 1000 10000 100000 1000000


Oc

cu

rre

nc

e (

%)

U

D

T

C

DM

Figure 8.23 Comparison of WW and DM for total nitrogen concentration in sewage sludge waste


101


0

5

10

15

20

25

30

35

40

1 10 100 1000 10000 100000 1000000

Total Phosphorus Concentration in mg/kg DM

Oc

cu

rre

nc

e (

%)

U

D

T

C

DM


0

5

10

15

20

25

30

35

40

45

50

1 10 100 1000 10000 100000 1000000


Oc

cu

rre

nc

e (

%)

U

D

T

C

DM

Figure 8.24 Comparison of WW and DM for total phosphorus concentration in sewage sludge waste


102


0

10

20

30

40

50

60

70

80

90

100

0.01 0.1 1 10 100 1000 10000

Lead Concentration in mg/kg DM

Oc

cu

rre

nc

e (

%)

U

D

T

C

DM


0

10

20

30

40

50

60

70

80

90

100

0.01 0.1 1 10 100 1000 10000


Oc

cu

rre

nc

e (

%)

U

D

T

C

DM

Figure 8.25 Comparison of WW and DM for lead concentration in sewage sludge waste


103

A8.5.3 Conclusion of the comparison between treatment type and difference between wet weight and dry matter

Considering wastes on a dry matter basis increases the relative concentration, as demonstrated in the above plots. However, the plots also seem to indicate that anaerobic digestion (both mesophilic and thermophilic) treatment processes produce higher concentrations when compared with composting treatment. Heat treatment seems also more likely to produce higher concentrations in the solid waste for all parameters.

However, there seems to be no significant changes of concentration between the different treatment types. This may be due to the high variability of the quality of the sludge itself and the difference in the quality of the sludge obtained from sewage treatment type.


104

A9. PATHOGENS

A9.1 Introduction

Pathogen data are not routinely collected for wastes applied to agricultural land, and therefore pathogen risk has been considered by conducting a review of the available literature.

Pathogens in organic wastes recycled to land present no agricultural benefit. Waste organic materials have the potential to contain a variety of pathogens, depending on the source of material. Pathogens pose an environmental risk for several reasons, including the transmission of human disease through the food chain, the transmission of diseases through farm animals and wildlife and the transmission of plant pathogens into the environment. There is also an issue of public perception which may override actual risks and require addressing.

Control of pathogen risks to human and animals can be controlled through sanitation and increasing the time between application and harvest, and washing and cooking of the crop by the consumer. The Safe Sludge Matrix (ADAS 2001) describes the usage of either partially sanitized (conventionaly treated) or extended sanitized (advanced treated) sewage sludge. The Animal By-products Order specifies the sanitation conditions required to treat animal by-products by either composting or anaerobic digestion processes which are integral in the PAS100 and PAS110 when animal by-products are included in the feedstock. It is important to remember that microbiological kill curves are exponential and that there is a risk of some pathogens surviving as resistant spores.

Pathogens are clearly a risk from materials of human and animal faecal origin. Plant pathogens may also however be a concern in infected material being composted and insufficiently sanitized and then used for land applications. For example there are concerns regarding the potential spread of plant pathogens such as sudden oak death through composted materials (Venglovsky et al. 2009). Composting of greenwaste by windrows may not be adequate for control of some pathogens. Composting by processes that match the requirements of the animal by-products order would pose less of a risk. The environmental risk from the spread of plant pathogens may be significant and not readily prevented or observed and therefore a precautionary approach is warranted.

A9.2 Animal by-products

The Animal By-products Regulations (ABPR) divide all wastes containing animal derived materials in to the following three categories:

Category 1: Very high risk (e.g. suspected BSE carcases).

Category 2: High risk (condemned meat; blood and gut contents).

Category 3: Low risk (catering waste from households, restaurants, former food, much slaughter house waste e.g. waste blood & feathers).

Category 1 wastes cannot be spread on land under any circumstances, and can only be disposed of by incineration or a plant approved to accept Category 1 materials.


105

Category 2 material may be spread to non-pasture land following processing to the method 1 standard (particle size <50 mm, heated to 133oC for 20 minutes, 3 bar pressure) set out in Annex V of Regulation (EC) 1774/2002. Optionally, this processing may be followed by composting or biogas treatment in a suitably approved plant prior to spreading. Certain category 2 materials (manure, digestive tract content, milk and colostrum) may be applied direct to land without treatment. They may only be applied to non-pasture land, except for manure4.

Category 3 wastes material may be spread to non-pasture land following treatment in a suitably approved processing, composting or biogas plant to the following standards.

Biogas plants

Maximum particle size 5 cm, minimum temperature 57oC for 5 hours.

Maximum particle size 12 cm, minimum temperature 70oC for 1 hour.

Closed reactor

Maximum particle size 40 cm, minimum temperature 60oC for 2 days.

Maximum Particle size 6 cm, minimum temperature 70oC for 1 hour.

Housed windrow

Particle size 40 cm, minimum temperature 60oC, minimum time at that temperature 8 days.

In addition, the Animal By-products Regulations stipulates that any plant for processing of organic materials must have clean and unclean areas, limiting any possibility of material by-passing the treatment process.

There has been extensive research to show that these requirements are enough to destroy most pathogens. A risk assessment conducted for Defra in 2000 (Gale et al. 2000), found that it is acceptable to apply these materials to land provided that the conditions stipulated in the ABPR are met.

The European Food Standards Agency (EFSA) are of the opinion (ESFA 2005) that where material is not pressure cooked (133oC, 20 minutes, 3 bar pressure) prior to composting or anaerobic digestion, there is some microbiological risk of contamination of final material. The current standard of 70oC for 60 minutes was not sufficient to inactivate all spore forming bacteria for Category 3 materials.

There remains the possibility of material by-passing the treatment process, by operators not following the requirements of the regulations. In the case of manures there is no requirement for treatment, and these can be spread directly on to pasture land. However, requirements for


106

time periods between application of the material and harvesting of crops are designed to mitigate these risks.

A9.3 Pathogens in compost

In the UK a significant proportion of compost is derived from source segregated green waste. This may also contain some food wastes which will then have been composted according to the ABPR. Plant pathogens are ubiquitous in the natural environment and most are indigenous to the UK and are part of the biological diversity of many habitats. For example, plant pathogens may be present within any plant material, whether in agricultural land forests, managed parkland, SSSIs, garden centres, plant nurseries and householder gardens. Other plant pathogens may have been accidentally introduced into the UK and become established in the wild to varying extents. Whatever the initial source, this background wild level of infection will exist. Consequently many plant pathogens may be expected to be present in green waste collected for composting.

There are many plant pathogens species, some of which are of greater concern than others. In this context the plant pathogen Phytopthora ramorum is notable as a plant pathogen attracting interest at national and international level. This plant pathogen is known to be infectious to several plant species and is a notifiable disease in the UK [UKSI n°2155 and Commission Decision (2002/757/EC)].

Chroni et al. (2009) reported that several experiments have been conducted to determine under which conditions composting is an effective process for the destruction of pathogens and pathogen indicators, such as Escherichia coli. This determination could be associated partly to maturity of compost (Finstein and Hogan 1993). Hess et al. (2004) concluded that multiple cycles of heating may be effective at destroying pathogens at lower temperatures than those needed in a single high temperature cycle. In contrast, Jiang et al. (2003) stated that slower heating up may increase heat resistance of pathogens during composting. In SSB, E. coli was detected in the raw material (8.37 * 104 CFU/g DW) and in the following five samples, despite the fact that temperature had reached 67°C by day 25 and only declined below detection limit after day 92. This was attributed to recontamination of the substrate in the centre of the pile by cooler material from its outer parts during turning. Re-contamination is common in windrow composting (Finstein and Hogan 1993; Haug 1993) and is an issue to consider when regulatory sanitisation requirements are based on specific time–temperature profiles. The total coliform population decreased during the process, but was detectable until the end. The Specific Oxygen Uptake Rate (SOUR) test was useful as an indicator of the advancement of the composting process and the stabilisation of the compost, verifying the value of respirometric tests for both compost stability assessments.

In 2003, WRAP commissioned a literature review (Jones and Martin 2003) on human and animal pathogens from green waste compost. This review found that whilst composting at a temperature of 55oC for at least 3 days was sufficient to eradicate most pathogens, resistant organisms such as C. perfringens, C. botulinum and the cysts and eggs of protozoan and helminth parasites may survive. There is also a risk that E. coli and salmonella may grow in the final compost product where the organic material has been poorly stabilised. The report concluded that more research was required to determine the risk of E. coli and salmonella developing in finished compost, and also on the survival of temperature resistant bacteria such as C. perfringens. E. coli and salmonella are the two surrogate pathogen parameters used to determine pathogen content in composts for the PAS100 quality compost standard.


107

An investigation in 2004 also commissioned by WRAP found that Enterohaemorrhagic E. coli, Salmonella typhimurium and S. enteriditis were not detectable after 1 hour at 50oC, and therefore concluded that green waste would not be a significant risk for the spread of most bacterial diseases. However, this study did not investigate the growth of these bacteria in the final compost product. The investigation also found that all but the most temperature resistant plant pathogens (Microdochiumnivale, Plasmodiophora brassicae and Tobacco Mosaic Virus, Microdochium nivale, Phytophthora nicotianae) were eradicated by temperatures of 52oC held for 7 days.

Composts applied under paragraph 7 exemptions are so-called „off specification composts‟ that do not meet the PAS100 standard for any reason. It may be considered that these materials are of higher risk than PAS100 quality composts, and therefore tighter controls placed on these materials regarding pathogens.

When considering the management of waste and the potential for dispersal of plant pathogens, it is clear that there are many pathways whereby plant pathogens may be transported to the receptor before the waste is treated in any composting plant. For example, routes of dispersal may occur if the infected waste material is home composted, transported in open trailers by householders to recycling centres, transported to transfer stations, and/or bulked up at transfer stations before being sent for treatment. Additionally there may be risks from the use of compost derived from infected waste that has not been adequately treated.

A literature review commissioned by WRAP in 2005 (Noble and Robert 2005) found that of the 60 pathogen and nermatode species investigated, compost temperature of 55oC is sufficient to eradicate all bacterial species. The fungal plant pathogens, Plasmodiophora brassicae, the causal agent of clubroot of Brassicas, and Fusarium oxysporum f. sp. lycopersici, the causal agent of tomato wilt, were more temperature tolerant. A compost temperature of at least 65°C for up to 21 days was required for eradication. Several plant viruses, particularly Tobacco Mosaic Virus (TMV) were temperature tolerant. However, there is evidence that TMV and Tomato Mosaic Virus are degraded over time in compost, even at temperatures below 50°C. The review singled out in-vessel composting systems as not being sufficient to degrade plant pathogens, as where the composting material is not turned, this increases the risk of cool-spots in the mix.

The risk assessment covered treated industrial wastes but did not include untreated waste that is currently spread to land, such as animal manures.

A9.4 Pathogens in sewage sludge

Sewage sludge spreading to agricultural land is controlled by the Sludge (use in Agriculture) 1989 Regulations.

The Sludge Regulations set various requirements. Which include:

minimum sampling and testing frequency of every 6 months;

set basic requirements for record keeping;

requires the farmer to work in sewage sludge soon after spreading.


108

The risk of pathogens in sewage sludge is controlled by the Safe Sludge Matrix in the UK. This is a voluntary code to which all water companies in the UK signed up to via Water UK. There are two levels of sludge treatment:

Conventional sludge treatment: Ensures that at least 99% (2 log reduction) of pathogens have been eradicated.

Enhanced Sludge Treatment: Ensures that 99.9999% (6 log reduction) of pathogens have been eradicated.

Which level of sludge treatment that the sludge is subjected to determines what end uses the sludge can be used for. Conventionally treated sludge cannot be used for crops intended for human consumption; they can only be used on animal feed and combinable crops. In contrast, enhanced treated sludge can be used on any type of crop, including salad crops intended for direct human consumption.

The principal method employed in the UK for pathogen inactivation is mesophilic anaerobic digestion. Mechanisms for pathogen inactivation in this process is still poorly understood. Smith (2008) also found that further research is necessary in to enteric microbes, and to quantify the decay of enteric viruses in sludge-treated agricultural soil. However, a report commissioned by UKWIR in 2003 found that the risk to the UK population is negligible where the Safe Sludge Matrix is adhered to. The Safe Sludge Matrix stipulates that a minimum of 10 month interval must be observed between spreading of the sludge and harvest, and this is sufficient to compensate for any inefficiencies in the sludge treatment process. Lang et al. (2003) found that thermally dried or composted sewage sludge did not contribute to background soil level of E Coli, but anaerobically digested sludges did increase e coli levels. However, the survival of the bacteria was limited to 3 months. A field trial carried out by ADAS also showed that E Coli O157, Salmonella and Campylobacter survived for up to 3 months in stored slurries, and Listeria for up to 6 months. They also found that the bacteria E Coli survived in the soil for up to 1 month after application.

Avery et al. found that sub-surface injection may reduce the risk of pathogen survival.

A9.5 Pathogens in manures

Defra‟s review of agriculture reviewed all Defra research in to agriculture between 1990 and 2005. This review found that there is significant pathogen risk to contaminate food crops (particularly ready to eat crops e.g. salads) where untreated animal manure is applied to the land. Pathogens of particular concern were bacteria salmonella, campylobacter and entero-hemorrhagic E Coli, viruses and parasitic protozoa such as Cryptosporidium and Giardia. These pathogens can also pose a risk in run-off to surface waters, resulting in illness in humans from bathing waters. The review also found that some pathogens could be transported up 1500 m from the site of spreading via aerosol dispersion, particularly from slurry spreading. The most cost effective barrier for transmission of these pathogens was longer storage periods prior to application, although this would increase ammonia, nitrous oxide and methane emissions during storage.

The Food Standards Agency issue guidance on spreading of manures on ready to eat crops. This guidance is non-statutory, but recommends that manure is either stored or composted prior to application, and a minimum of 12 months between spreading and harvesting.


109

A9.6 Pathogens in paper sludges

Paper sludges are by weight one of the largest materials currently spread to agricultural land in the UK. Pathogens are not thought to pose any issue of significant environmental concern with paper mill sludges (WESA and Bates 2002, SEPA 1999).


110

A10. KEY FINDINGS AND CONCLUSIONS

A10.1 Comments on general approach

In this study, an approach of normalising the application rates of different wastes to agricultural land has been used, based on either applying the wastes to provide either the same PAN as a nitrogen fertiliser or the same increase in stable SOC if applied as a soil conditioner. This has allowed a comparison between wastes for many other parameters such as nutrient addition (P, K, N and S), heavy metal additions to soil, and gaseous emissions such as NH3 CO2 and N2O.

In this study the methodology has been applied to the mean composition of different wastes types and waste treatments although this has in some cases required the use of assumed values in order to demonstrate the value of the approach. The waste characterisation data collected in this study was not complete for all the wastes. However, the data indicates that there may be a wide variation in any particular parameter for the same type of waste and therefore the mean values used here should only be considered as demonstrating the approach. Applying the methodology to specific wastes would provide a comparative judgement on its potential agronomic benefits and environmental impacts. This comparison may then be used to limit the application rates for the particular waste to control environmental risks.

The waste characteristics used in this assessment included many that are routinely measured in the laboratory on organic wastes, e.g. for PAS100 characterisation of composts. Additionally the behaviour of the waste when applied to soil with respect to certain waste characteristics is not known. Such behaviour may be critical for obtaining a greater understanding of the agronomic and environmental risks associated from any particular waste applied to land. To achieve such an understanding it may be necessary to carry out soil tests and relate the results of these to laboratory waste characterisation tests. Given the diverse range of soil types these characterisations may need to cover the range of soil types available. It may then be possible to develop correlations between appropriate laboratory waste characterisation tests and behaviour of the waste in the soil environment. There may also be a need to undertake additional research to elucidate some particular relationships between laboratory test and behaviour in soils.

This level of monitoring may be criticised as being excessive in many cases. However the approach could be accommodated within a traditional tiered risk assessment where initial screening using relatively simple waste characterisation data may indicate that for some specific waste parameters the risks may be of concern and more detailed data is required to assess the risks fully.

A10.2 Agronomic benefit

Agronomic benefit from the application of organic wastes to soil may be derived from the plant nutrient (fertilizer) content of the wastes or from general soil improvement. Organic wastes may vary considerably in these properties. Nutrients in organic wastes need to be in an available form or readily decomposed into an available form. For example anaerobic digestates and manures are a good source of N whereas stabilised composts are not.


111

However composts may be good sources of other nutrients such as P, K, Mg and S and it is possible that mixing organic wastes with different nutrient properties may allow optimisation of organic applications for maximum agronomic benefit.

It is important that the loadings of available nutrients match the crop requirement to minimise the risk of excess nutrients posing a threat as pollutants. Therefore it is also important that the wastes are characterised using suitable laboratory tests that provide a good prediction of the availability of the nutrient in the soil environment. This is particularly so for major plant nutrients such as N and P. This study indicates that such tests are rarely carried out and a full understanding of the waste crop nutrient is not available. It is suggested that more detailed waste testing should be carried out (see section A9.7) to provide such knowledge.

The view is taken that it is not clear how much P is available in different organic wastes and that it is possible that individual wastes will vary significantly depending on the exact source and treatment of the waste. An appropriate waste characterisation method is not available. It is premature to provide assurance that waste characterisation based on Olsen-P or similar provides a reliable prediction of the P risks. A conservative view is taken and it is assumed that a significant fraction of the total P is available.

The crop nutrient guidance provided in RB209 is a good model of how such issues may be controlled. Appropriate values for available N, P, K, Mg and S need to be developed for other organic wastes. However given the wide variation in characteristics of wastes of the same type it may be better to describe the tests required and the application rates and not describe compositional data for wastes.

A10.3 Soil quality issues

Soil quality is a complex area with many facets that could not be accommodated in the scope of this study. In general the application of stabilised organic matter to soil is accepted as good practice benefiting the soil in many ways such as improving water retention, and soil structure thereby providing a better environment for soil organisms and plants.

However the characteristics of the waste need to be understood and many of these important waste properties may not be routinely monitored. For example a key parameter is the degree of stabilisation and the maximum biological oxygen demand of the waste. This would impact on the short term aerobicity of the soil and the ultimate extent of carbon sequestration in the soil. Other important factors are the accumulation of metals in soils. This study indicates that whatever waste is applied they will apply metals to the soil and if applied at high loadings may increase soil metal contents significantly. The soil metal limits applied in the regulation of sewage sludge provide a good approach to controlling these issues.

The biodegradability of organic wastes is believed to be an important factor with respect to the impact on soil biology and physico-chemical structure. Characterisation of wastes for biodegradability is rare and the impact of this on its behaviour in soil is also rare. It is recommended that further studies are carried out to provide a better understanding on what waste characteristics may be used to predict the impact of the waste in the soil environment.

Risks from organic pollutants present in the wastes are also of concern. This is a complex issue as there are many potential organic pollutants and monitoring all would be a costly approach The use of biological screening tests may be more cost effective approach. These could be for plant toxicity, microtox, endocrine disruption and other simple ecotoxicological


112

tests. Backing up the tests with a less frequent screen for selected organic pollutants would build up a databank for any particular waste.

Alternatively the use of soil limits for organic pollutants (mimicking soil metal limits) may be applied as described here. Organic pollutants in soil may be subject to microbial decomposition which may need to be taken into account.

A10.4 Greenhouse gas emissions

As microbial activities in soils are a key component of the greenhouse gas emissions and mitigation it is important that the impact of organic waste applications to soil are understood in order to minimise impacts. This is an area of immense study, particularly with respect to N2O emissions and CH4 consumption in soils. In this study a simple model for the estimation of N2O emissions based on total N loadings has been used, as more comprehensive data was not available. This may be seen as being too simplistic as emissions will depend on many soil, weather and waste characteristics. This study indicates how this method may be used as an initial comparative screen. Further work is recommended to define a relationship between measured waste characteristics such as biodegradability, mineral N, and mineralisable N that may then be applied to provide a more reliable prediction.

Estimating CH4 uptake or emission from soil is even more difficult because it depends on the biology of anaerobic methanogenic and aerobic methane oxidising bacteria. These activities will be governed in part by the soil conditions and the biodegradability of the waste. However it is important that the CH4 oxidising microbes are not adversely affected by other components of the waste and further work is recommended to review the current literature to determine if a predictive model based on suitable waste characteristics may be developed.

A10.5 Ground and surface water

In the absence of laboratory leaching test data, theoretical leachate concentrations for selected parameters were calculated by Consim© from total composition data on a dry weight basis. In the Tier 1 risk assessment the calculated leachate concentrations were then compared with relevant water quality standards and in the case of groundwater published background water quality in the UK. (Shand et al. 2007). To provide a worst case assessment the 95%ile leachate concentration obtained for each waste category was used to assess compliance or breach of water quality benchmarks. It should be noted that the Level 1 approach and use of essentially predicted maximum concentrations provides a worst case assessment of risk. The assessment has been carried out for priority hazardous substances (formally List I substances) which must not be discharged to groundwater under Directive 2008/105/EC, and priority substances (formally List II substances) for which discharge to groundwater should be limited. A more comprehensive report of the assessment process for surface and ground water is provided in Appendix B.

Priority hazardous substances Cd and Hg (formerly List I substances) are present in almost all of the wastes reviewed for this project. However, for the most part these are generally present at concentrations that would require little or no dilution to be below background water quality benchmarks. The presence of these substances is therefore not considered to present a significant risk to the environment (surface and ground waters). The exceptions, where


113

calculated leachate concentrations were much higher than background concentrations or environmental quality standards for surface or groundwater were some datasets for sewage sludge (waste type 20) and the organic fraction of municipal solid waste (waste type 60).

From the assessment of priority substances only Pb and Ni are higher than background concentrations (10 to 20 times higher) for sewage sludge and organic fraction of MSW. Leachate produced by application of these wastes to land, may pose problems for the environment. Calculated chromium and copper leachate concentrations for almost all wastes reviewed exceed background concentrations by 2500 and 5 times and may therefore pose a possible significant risk to surface and groundwater quality. The exceptions are waste from processing of food and „other‟ categories. Calculated leachate concentrations for zinc are not deemed to present a risk to surface and groundwater. Possible mitigation could include control of the input waste streams to the treatment processes, blending of the outputs, a reduction in the quantity of waste applied to land and good agricultural practice related to applications in the proximity of surface water. This requires routine monitoring of the treatment residues for total and leachable Pb, Ni, Cr, Cd, Cu and Zn which are not being undertaken at present.

In many areas of England and Wales phosphorus and nitrogen levels are high, and many water bodies are failing to achieve good status. Using the 95th percentile concentrations, most of the wastes studied for this project have the potential to cause harm to surface and ground waters due to high concentrations of nitrogen and phosphorus. The exception to this statement relate to phosphorus levels in the „other‟ waste category. The application of wastes that contain high levels of phosphorus and nitrogen to land should be undertaken with care to avoid compromising ability to achieve good status for water bodies in England and Wales.

In the absence of other relevant benchmarks potassium concentrations were compared to groundwater background concentrations for the UK and indicate that predicted concentrations are close to background levels, and therefore not deemed to present a risk to surface or groundwater.

It should be noted that for this study only a Level 1 risk assessment was undertaken due to paucity of data, and variability in environmental conditions across England and Wales. Leaching concentrations were calculated from total concentrations obtained in the data collation. As such the conclusions represent a worst case scenario, but indicate where extra care may be needed. It is recommended that a robust leachability dataset is compiled in future research on these materials to allow a comprehensive risk assessment to be undertaken in the future. Mitigation factors as outlined in Defra‟s Code of Good Agricultural Practice are very important for the protection of surface water and groundwater and should be followed.

A10.6 Pathogen risks

Pathogens to plants and animals may be present in all forms of organic wastes. The risks from these are largely being addressed by the wastes undergoing pre-treatments such as heat treatment, composting and AD. Very few organic wastes are envisaged being applied to soil untreated.

However pathogens are capable of growth and reproduction, and mutation to more virulent and drug resistant strains. Hence it is important that a constant vigilance is kept of the risks from pathogens,


114

A10.7 Suggested waste characterisation required

The suggested waste characteristics and testing that would be required in order to provide a greater understanding of the agronomic benefit and environmental impact are given in Table 9.1.

Table 10-1 Suggested waste characterisation testing required to improve prediction of agronomic benefit and environmental impact of organic wastes spread to land

Waste Parameter

Benefit or adverse impact

Relevant waste tests Correlating soil test

NH3 Air emission Extractable NH3 pH Dry matter

Testing to confirm correlations used for manure applications can be applied to all wastes

PAN Plant nutrient content

Laboratory soil N mineralization test

Confirm reflects soil conditions.

Available P Plant nutrient content. P mobility as pollutant

Laboratory P availability testing.

Confirm reflects P availability in soil.

Biodegradability Soil quality, nutrient releases, air emissions

Short and long term aerobic biodegradability tests to determine initial rate of degradation and the full extent of degradation

Confirm and correlate with waste decomposition in soil

Organic pollutants

Soil contamination, crop contamination and diffuse pollution

Possible ecotoxicological screening tests

Correlation with soil fate and effect studies


115

REFERENCES

Abdullahi, Y.A., Akunna, J.C. White, N.A., Hallett, P.D. and Wheatley, R. (2008). Investigating the effects of anaerobic and aerobic post-treatment on quality and stability of organic fraction os municipal solid waste as soil amendment. Bioresource Technol. 99, p8631.

Adani, F., Ricca, G. (2004). The contribution of alkali soluble (humic acid-like) and unhydrolyzed-alkali soluble (core-humic acid-like) fractions extracted from maize plant to the formation of soil humic acid. Chemosphere 56, 13–22.

ADAS (2007). Biowaste materials: Quality and suitability for land recycling. EA project PO 30129085. Draft final report.

ADAS (2001) Guidelines for the application of sewage sludge to agricultural land, The Safe Sludge Matrix, British Retail Consortium, Water UK and ADAS, April 2001. www.adas.co.uk

AFOR (2009). Be alert to herbicide active ingredients aminopyralid and chloropralid. www.organics-recycling.org.uk/index.php?view+article&catid=1%3Alatest-news&...

Aggelides, S.M., Londra, P.A. (2000). Effects of compost produced from town wastes and sewage sludge on the physical properties of a loamy and a clay soil. Bioresour. Technol. 71, 253–259.

Amon, B., Kryvoruchko, V., Amon, T., Zechmeister-Boltenstern, S. (2006). Methane, nitrous oxide, and ammonia emissions during storage and after application of dairy cattle slurry and influence of slurry treatment. Agriculture, Ecosystems and Environment 112, 153–162.

Barral M. T., Paradelo R., Moldes A. B., Dominguez M. and Diaz-Fierros F. (2009). Utilization of MSW compost for organic matter conservation in agricultural soils of NW Spain. Resources, Conservation and Recycling (in press).

Battaglia, A., Calace, N., Nardi, E., Petronio, B., Pietroletti, M. (2007). Reduction of Pb and Zn bioavailable forms in metal polluted soils due to paper mill sludge addition. Effects on Pb and Zn transferability to barley. Bioresource Technol. 98, 2993–2999.

Bernal, M., Navarro, A.F., Roig, A., Cegarra, J., Garcia, D. (1996). Carbon and nitrogen transformation during composting of sweet sorghum bagasse. Biology and Fertility of Soils 22, 141±148.

Bernal, M.P., Sanchez-Monedero, M.A., Pareddes, C. and Roig, A. (1998). Carbon mineralization from organic wastes at different composting stages during their incubation with soil.

Bernet, N. and Beline, F. (2009). Challenges and innovations on biological treatment of livestock effluents. Bioresource Technol. (in press).

Bishop, P.L., Godfrey, C. (1983). Nitrogen variations during sludge composting. BioCycle 24, 34-39.

Bhogal, A., Nicholson, F.A., Chambers, B.J. (2009). Organic carbon additions: effects on soil bio-physical and physico-chemical properties. European Journal of Soil Science 60, 276-286.

http://www.organics-recycling.org.uk/index.php?view+article&catid=1%3Alatest-news&


116

Bouwman, A.F. (1996). Direct emission of nitrous oxide from agricultural soils. Nutrient Cycling in Agroecosystems 46, 53-70.

BSFP (2008). British Survey of Fertiliser Practice. https//statistics.defra.gov.uk/esg/bsfp.htm.

BS EN 13650. Soil improvers and growing media – extraction of aqua regia soluble elements.

BS EN 13651. Soil improvers and growing media - Extraction of calcium chloride/DTPA (CAT) soluble elements.

BS EN 13652. Soil improvers and growing media – Extraction of water soluble nutrients and elements.

Burger, M., Venterea, R. (2008). Nitrogen Immobilization and Mineralization Kinetics of Cattle, Hog, and Turkey Manure Applied to Soil. Soil Science Society of America Journal 72, 1570-1579.

Burton, C.H. (2009). Reconciling the new demands for food protection with environmental needs in the management of livestock wastes. Bioresource Technol. (in press).

Busby, R.R., Torbert, H.A. and Gebhart, D.L. (2007). Carbon and nitrogen mineralization of non-composted and composted municipal solid waste in sandy soils. Soil Biol. Biochem., 39, p1277.

Cabrera, M.L., Kissel, D.E., Vigil, M.F. (2005). Nitrogen mineralization from organic residues: research opportunities. Journal of Environmental Quality 34, 75–79.

Calace, N., Campasi, T., Lancondini, A., Leoni, M., Petronio, B., Pietroletti, M. (2005). Metal-contaminated soil remediation by means of paper mill sludges addition: chemical and ecotoxicological evaluation. Environ. Pollut. 136 (3), 485–492.

Carpenter, F., Ferna´ndez, I.J. (2000). Pulp sludge as a component in manufactured topsoil. J. Environ. Qual. 29, 387–397.

Chambers, B. J. et al. (2003). Effects of sewage sludge applications on agricultural soils on sol microbial activity and the implications for agricultural productivity and long-term soil fertility.

Chantigny, M.H., Angers, D.A., Rochette, P., Bélanger, G., Massé, D. (2007). Gaseous nitrogen emissions and forage nitrogen uptake on soils fertilized with raw and treated swine manure. Journal of Environmental Quality 36, 1864-1872.

Chantigny, M.H., Rochette, P., Angers, D.A., Massé, D., Côté, D. (2004). Ammonia volatilization and selected soil characteristics following application of anaerobically digested pig slurry. Soil Science Society of America Journal 68, 306–312.

Chroni, C., Kyriacou, A., Manios, T., Lasaridi, K-E. (2009). Investigation of the microbial community structure and activity as indicators of compost stability and composting process evolution. Bioresource Technology 100 3745–3750.

Clemens, J., Trimborn, M., Weiland, P., Amon, B. (2006). Mitigation of greenhouse gas emissions by anaerobic digestion of cattle slurry. Agriculture, Ecosystems and Environment 112, 171–177.


117

Comans, van der Sloot, Bpnpuvrie (1993), Geochemical reactions controlling solubility of major and trace elements during leaching of municipal solid waste incinerator residues, Municipal Waste Combustion Conference, Williamsburg V.A.

Conrad, R. (1996). Soil microorganisms as controllers of atmospheric trace gases (H2, CO, CH4, OCS, N2O and NO). Microbiol. Rev., 60, p609.

Courtney, R.G., Mullen, G.J. (2008). Soil quality and barley growth as influenced by the land application of two compost types. Bioresource Technology 99, 2913–2918.

de Bertoldi, M., Citernesi, U., Gricelli, M. (1980). Bulking agents in sludge composting. Compost Science and Land Utilization 21 (1), 32-35.

de Bertoldi, M., Vallini, G., Pera, A. (1985). Technological aspects of composting including modelling and microbiology. In: Gasser, J.K.R (Ed.), Composting of Agricultural and Other Wastes. Elsevier Applied Science Publishers, New York, USA, pp. 27-41.

De Boer, H.C. (2008). Co-digestion of animal slurry can increase short-term nitrogen recovery by crops. Journal of Environmental Quality 37, 1968-1973.

Defra (2002). Ammonia in the UK. Defra publication PB0000, http://www.defra.gov.uk.

Defra (2003). Code of Practice for the Agricultural Use of Sewage Sludge. http://www.defra.gov.uk.

DEFRA Research In Agriculture And Environmental Protection Between 1990 And 2005: Summary And Analysis (ES0127)

Domene, X., Alcaniz, J.M. and Andres, P. (2007). Ecotoxicological assessment of organic wastes using the soil collembolan Folsomia candida. Applied Soil Ecology, 35, p461.

EA (2008). Bioaerosols in waste composting: deriving source terms and characterising profiles

EA (2009). Quality Protocol Anaerobic Digestate. http://www.environment-agency.gov.uk/static/documents/business/090814_AD_QP_v32_FINAL.pdf

EA (2009b). Contaminated Land Exposure Assessment (CJEA). www.environment-agency.gov.uk/research/planning/33714.aspx.

Eghball, B., Gilley, J.E. (1999). Phosphorus and nitrogen in runoff following beef cattle manure or compost application. J. Environ. Qual. 28, 1201–1210.

Eghball, B., Power, J.F. (1999). Composted and noncomposted manure application to conventional and no-tillage systems: corn yield and nitrogen uptake. Agron. J. 91, 819–825.

Eghball, B. (2003). Leaching of phosphate fractions following manure or compost application. Commun. Soil Sci. Plant. Anal., 34, p2803-2815.

Elliott, H.A. and O‟Conner, G.A. (2007). Phosphorous management for sustainable biosolid recycling in the United States. Soil Biol. Biochem., 39, p1318.

http://www.defra.gov.uk/


http://www.environment-agency.gov.uk/static/documents/business/090814_AD_QP_v32_FINAL.pdf

http://www.environment-agency.gov.uk/static/documents/business/090814_AD_QP_v32_FINAL.pdf


118

Elliot, H.A., O´Connor, G.A., Brinton, S. (2002). Phosphorous leaching from biosolids amended sandy soils. Journal of Environmental Quality 31, 681–689.

Epstein, E. (1997). The Science of Composting. Technomic Publishing Co., Inc., Lancaster, PA, USA.

The EFSA Journal (2005) 264, 1-21. Opinion on the safety vis -à-vis biological risks of biogas and compost treatment standards of animal by-products.

Esteller, M.V., Matrinez-Valides, H., Garrido, S. and Uribe, Q. (2009). Nitrate and phosphate leacing in a phaeozem soil treated with biosolids, composted biosolids and inorganic fertilizers. Waste Management, 29, p1936-1944.

Estrada, I.B., Gomez, E., Aller, A. and Moran, A. (2007). Microbial monitoring of the influence of the stabilization degree of sludge when applied to soil. Bioresource Technology, 97, p1308.

European Commission (Directorate-General Environment ENV A.2). (2001). Biological treatment of biowaste, working document, 2nd draft. Available from: <http://www.compost.it/www/pubblicazioni_on_line/biod.pdf> (verified 19.09.08).

Evanylo, G., Sherony, C., Spargo, J., Starner, D., Brosius, M., Haering, K. (2008). Soil and water environmental effects of fertilizer-, manure-, and compost-based fertility practices in an organic vegetable cropping system Agriculture, Ecosystems and Environment 127, 50–58.

Fabrizio, A., Tambone, F., Genevini, P. (2009). Effect of compost application rate on carbon degradation and retention in soils. Waste Management 29, 174–179.

Feldkinchner, D., Wang, C., Gower, S., Kruger, E., Ferris, J. (2003). Effects of nutrient and paper mill biosolids amendment on growth and nutrient status of hardwood forest. Forest Ecol. Manag. 177, 95–116.

Finstein, M.S., Hogan, J.A. (1993). Integration of composting process microbiology, facility structure and design-making. In: Hoitink, HAJ, Keener HM, (Eds.), Science and Engineering of Composting. Renaissance Publications, Ohio, pp. 1–23.

Fjallborg, B., Ahlberg, G., Nilsson, E. and Dave, G. (2005). Identification of metal toxicity in sewage sludge leachate. Env. Int. 31, p25.

Flessa, H., Dörsch, P., Beese, F. (1995). Seasonal variation of N2O and CH4 fluxes in differently managed arable soils in southern Germany. Journal of Geophysical Research 100 (D11), 23114-23124.

Foley, B., Cooperband, L. (2002). Paper mill residuals and compost effects on soil carbon and physical properties. J. Environ. Qual. 31, 2086–2095.

Francou, C., Poitrenaud, M., Houot, S. (2005). Stabilization of organic matter during composting: influence of process and feedstocks. Compost Science & Utilization 13, 72–83.

Frossard, E., Skrabal, P., Sinaj, S., Bangerter, F. and Traore, O. (2002). Forms and exchangeability of inorganic phosphate in composted solid organic wastes. Nutr. Cycl. Agroecosys. 62, p103-113


119

FSA (2009) Managing Farm Manures for Food Safety Guidelines for growers to reduce the risks of microbiological contamination of ready-to-eat crops

Gabrielle, B., Da-Silveira, J., Houot, S., Francou, C. (2004). Simulating urban waste compost effects on carbon and nitrogen dynamics using a biochemical index. Journal of Environmental Quality 33, 2333–2342.

Gagnon, B., Lalande, R., Fahmy. (2001). Organic matter and aggregation in a degraded potato soil as effected by raw and composted pulp residue. Biology and Fertility of Soils 34, 441–447.

Gale et al. (2000), Risk Assessment: Use of Composting and Biogas Treatment to Dispose of Catering Waste Containing Meat (WRc)Gerritse, R.G., Vriesema, R. (1984). Phosphate distribution in animal waste slurries. Journal of Agricultural Science 102, 159–161.

Godley, A., Lewin, K., Frederickson, J., Smith, R. and Blakey, N. (2007). Application of DR4 and BM100 biodegradability tests to treated and untreated organic wastes. Sardinia

Godley, A., Frederickson, J., Turrell, J., Smith, R. and Blakey, N. (2008). Characterisation of biodegradable residues from MBT treatment of MSW. Conference proceedings Waste 2008. Stratford-Upon-Avon, 16-17 Sept.

Granli, T., Bøckman, O.C. (1994). Nitrous oxide from agriculture. Norwegian Journal of Agricultural Science, Supplement No. 12. 128 pp.

Güngör, K., Karthikeyan, K.G. (2008). Phosphorus forms and extractability in dairy manure: A case study for Wisconsin on-farm anaerobic digesters. Bioresource Technology 99,425–436.

Hansen, M.N., Kai, P., Moller, H.B. (2006). Effects of anaerobic digestion and separation of pig slurry on odor emission. Applied Engineering in Agriculture 22, 135–139.

Haug, R.T. (1993). The Practical Book of Compost Engineering. Lewis Publishers, Boca Raton.

Hénault, C., Devis, X., Lucas, J.L., Germon, J.C. (1998). Influence of different agricultural practices (type of crop, form of N-fertilizer) on soil nitrous oxide emissions. Biology and Fertility of Soils 27, 299-306.

Hess, T.F., Grdzelishvili, I., Sheng, H., Hovde, C.J. (2004). Heat inactivation of E. coli during manure composting. Compost Sci. Util. 12 (4), 314–322.

Higgins, S.F., Coyne, M.S., Shearer, S.A. and Crutchfield, J.D. (2005). Determining nitrogen fractions in swine slurry. Bioresource Technology. 96, 1081-1088.

Hjorth, M., Nielsen, A.M., Nyord, T., Hansen, M.N., Nissen, P., Sommer, S.G. (2009). Nutrient value, odour emission and energy production of manure as influenced by anaerobic digestion and separation. Agronomy for Sustainable Development 29, 329–338.

Hue, N.V., Ikawa, H., Silva, J.A. 1994. Increasing plant available phosphorus in an ultisol with a yard waste compost. Communications in Soil Science and Plant Analysis 25, 3291–3303.

IAEUKA (2007). Inventory of Ammonia Emissions in UK Agriculture. Htttp://www.northwyke.bbsrc.ac.uk/AmmoniainventoryWebsite/Ammoniainventory.html.


120

Iglesias-Jimenez, E., Alvarez, C.E. 1993. Apparent availability of nitrogen in composted municipal refuse. Biology and Fertility of Soils 16, 313–318.

Illera, V., Walter, I., Souza, P. and Cala, V. (2000). Short-term effects of biosolid and municipal solid waste applications on heavy metal distribution in a degraded soil under a semi-arid environment. The Science of the Total Environ., 255, p29-44.

IPCC. (2006). IPCC Guidelines for National Greenhouse Gas Inventories. Chapter 11: N2O Emissions from Managed Soils, and CO2 Emissions from Lime and Urea Application. 54 pp.

Jenkinson, D.S. (1988). Soil organic matter and its dynamics. In: Russels's SOil Conditions and Plant Growth (ed. A Wild, pp 4564-607. Longmans, Harlow.

Jiang, X., J. Morgan and M.P. Doyle. (2003) Fate of Escherichia coli O157:H7 during Composting of Bovine Manure in a Laboratory-Scale Bioreactor. Journal of food protection. 66(1):25-30.

Jones and Martin (2003) A Review Of The Literature On The Occurrence And Survival Of Pathogens Of Animals And Humans In Green Compost, Institute For Animal Health

Kaiser, E.A., Kohrs, K., Kücke, M., Schnug, E., Munch, J.C., Heinemeyer, O. (1997). Importance of increasing N-fertilisation levels for the N2O-emissions from different arable crops. In: Proceedings of the 7th International Workshop on nitrous oxide emissions, Cologne, 21-23 April 1997. pp 171-176

Keeling, A.A., McCallum, K.R., Beckwith, C.P. (2003). Mature green waste compost enhances growth and nitrogen uptake in wheat (Triticum aestivum L.) and oilseed rape (Brassica napus L.) through the action of water-extractable factors. Bioresour. Technol. 90, 127–132.

Kirchmann, H., Lundvall, A. (1993). Relationship between N immobilization and volatile fatty acids in soil after application of pig and cattle slurry. Biology and Fertility of Soils 15, 161-164. reported N immobilization in soil only half that from digested compared with undigested slurry, due to less easily-decomposable C in the digested slurry.

Komilis, D.P., Tziouvaras, I.S. (2009). A statistical analysis to assess the maturity and stability of six composts Waste Management 29, 1504–1513.

Körschens, M., Müller, A. (1996). The static experiment Bad Lauchstädt, Germany. In: Powlson DS, Smith P, Smith JU, (Eds.), Evaluation of Soil Organic Matter Models using Existing, Long-Term Datasets, NATO ASI Series I, vol. 38. Springer, Heidelberg, Germany, pp. 369–376.

Kupper, T., Bucheli, T.D. Brandli, R.C., Ortelli, D., Edder, P. (2008). Dissipation of pesticides during composting and anaerobic digestion of source-separated organic waste at full-scale plants. Bioresource Technol., 99, p7988.

Lang, N.L., Smith, S.R., Bellett-Travers, D.M. and Pike, E.B. (2004) Pathogen decay in soil – the final barrier to enteric disease transmission. Sustainable Land Application Conference, 4 – 8 Jan, Orlando, FloridaLarney, F.J., Hao, X. (2008). A review of composting as a management alternative for beef cattle feedlot manure in southern Alberta, Canada. Bioresource Technol. 98, 3221–3227.


121

Levy, J.S., Taylor, B.R. (2003). Effects of pulp mill solids and three composts on early growth of tomatoes. Bioresource Technol. 89 (3), 297–305.

Litterick, A.M., Harrier, L., Wallace, P., Watson, C.A., Wood, M. (2004). The role of uncomposted materials, composts, manures, and compost extracts in reducing pest and disease incidence and severity in sustainable temperate agricultural and horticultural crop production: a review. Crit. Rev. Plant Sci. 23, 453–479.

Madrid, F., López, R., Cabrera, R. (2007). Metal accumulation in soil after application of municipal solid waste compost under intensive farming conditions. Agr. Ecosys. Environ. 199, 249–256.

Maeda, M., Zhao, B., Ozaki, Y., Yoneyama, T. (2003). Nitrate leaching in an Andisol treated with different types of fertilizers. Environmental Pollution 121, 477–487.MAFF (1986). The analysis of agricultural materials. Reference book 427 3rd Edition. ISBN 0 11 242762 6.

MAFF. (1986). The Analysis of Agricultural Materials. Reference Book 427. HMSO, London.

MAFF (2000). Fertiliser Recommendations for agricultural and horticultural crops (RB209), 7th Edition. www.defra.gov.uk. ISBN 0 11 243058 9.

Mamo, M., Rosen, C.J., Halbach, T.R. (1998). Nitrogen availability and leaching from soil amended with municipal solid waste compost. J. Environ. Qual. 28, 1074–1082.

Mamo, M., Rosen, C.J., Halbach, T.R., Moncrief, J.F. (1999). Corn yield and nitrogen uptake in sandy soils amended with municipal solid waste compost. J. Prod. Agric. 11, 469–475.

Marcato, C.E., Pinelli, E., Pouech, P., Winterton, P. and Guiress, M. (2008). Partical size and metal distribution in anaerobically digested pig slurry.

Marcato, C-E., Mohtar, R., Revel, J-C., Pouech, P., Hafidi, M. and Guiresse, M. (2009). Impact of anaerobic digestion on organic matter quality in pig slurry. Int. Biodeterioration and Biodegradation, 63, p260.

Martens, D.A. (2000). Plant residue biochemistry regulates soil carbon cycling and carbon sequestration. Soil Biology and Biochemistry 32, 361–369.

Meijide, A., Garcia-torres, L., Arce, A. and Vallejo, A. (2009). Nitrogen oxide emissions affected by organic fertilisation in a non-irrigated Mediterranean barley field. Agri. Ecosystems and Env., 132, p106.

Miaomioa, H., Wenhong, L., Xinqiang, L., Donglei, W. and Guangming, T. (2009). Effect of composting process on phytotoxicity and speciation of copper, zinc and lead in sewage sludge and swine manure. Waste Management, 29, p590.

Mikutta, R., Kleber, M., Torn, M.S., Jahn, R. (2006). Stabilization of soil organic matter: association with minerals or chemical recalcitrance? Biogeochemistry 77, 25–56.

Minami, K. and Takat, K. (1997). Atmospheric methane: sources, sinks, and strategies for reducing agricultural emissions. Wat. Sci. Tech., 36, p509.



122

Mkhabela, M.S., Warman, P.R. (2005). The influence of municipal solid waste compost on yield, soil phosphorus availability and uptake by two vegetable crops grown in a Pugwash sandy loam soil in Nova Scotia. Agriculture, Ecosystems and Environment 106, 57–67.

Moller, K. and Stinner, W. Effects of different manuring systems with and without biogas digestion on soil mineral nitrogen content and on gaseous nitrogen losses (ammonia, nitrous oxides). Europ. J. Agronomy, 30, p1.

Mondini, C., Cayuela, M.C., Sinicco, T., Sanchez-Mondero, M.A., Bertolone, E. and Bardi, L. (2008). Soil application of meat and bone meal. Short-term effects on mineralization dynamics and soil biochemical and microbiological properties. Soil Biol. Biochem., 40, 462-474.

Moral, R., Paredes, C., Bustamante, M.A., Marhuenda-Egea, F., Bernal, M.P. (2009). Utilisation of manure composts by high-value crops: Safety and environmental challenges. Bioresource Technology in press.

Müller, C., Kammann, C., Burger, S., Ottow, J.C.G., Grünhage, L., Jäger, H.-J. (1997). Nitrous oxide emissions from frozen grassland soil and during thawing. in: Becker, K.H. & Wiesen, P. (eds.): Proc. 7th International Workshop on Nitrous Oxide Emissions, Cologne, Germany, April 21-23 1997. Berichte der Physikalischen Chemie (Bergische Universität, Gesamthochschule Wuppertal) 41, 327-335.

Nicholson, F.A., Chambers, B.J., Williams, J.R. and Unwin, R.J. (1999). Heavy metal contents of livestock feeds and animal manures in England and Wales. Bioresource Technol., 70, p23.

Noble and S.J. Roberts (2005) A Review Of The Literature On Eradication Of Plant Pathogens And Nematodes During Composting, Disease Suppression And Detection Of Plant Pathogens In Compost, Horticulture Research International

Nunes, J.R., Cabral, F., López-Pineiro, A. (2008). Short-term effects on soil properties and wheat production from secondary paper sludge application on two Mediterranean agricultural soils. Bioresource Technology 99, 4935–4942

Nyberg, K., Sundh, I., Johansson, M., Schnürer, A. (2004). Presence of potential ammonia oxidation (PAO) inhibiting substances in anaerobic digestion residues. Applied Soil Ecology 26, 107–112.

Oades, J.M. (1993). The role of biology in the formation, stabilization and degradation of soil structure. Geoderma 56, 377–400.

Oehl, F., Frossard, E., Fliessbach, A., Dubios, D. and Oberson, A. (2004). Basal organic phosphorous mineralization in soils under different farming systems. Soil boil. Biochem., 36, p 667.

Paavola, T. and Rintala, J. (2008). Effects of storage on characteristics and hygienic quality of digestates from four co-digestion concepts of manure and biowaste. Bioresource technol. 99, p7041.

Pain, B.F., Misselbrook, T.H., Clarkson, C.R., Rees, Y.J. (1990). Odour and ammonia emissions following the spreading of anaerobically-digested pig slurry on grassland. Biological Waste 34, 259–267.


123

Paustian, K., Andrén, O., Janzen, H.H. (1997). Agricultural soils as a sink to mitigate CO2 emission. Soil Use and Management 13, 230–244.

PAS100:2005. Specification for compost.

Peralta-Videa, J.R., Lopez, M.L., Narayan, M., Saupe, G. and Gardea-Torresdey, J. (2009). The biochemistry of environmental heavy metal uptake by plants: implications for the food chain.

Perucci, P. (1992). Enzyme activity and microbial biomass in a field soil amended with municipal refuse. Biology and Fertility of Soils 14, 54–60.

Petersen, S.O. (1999). Nitrous oxide emissions from manure and inorganic fertilizers applied to spring barley. Journal of Environmental Quality 28, 1610–1618.

Pognani, M., D‟Imporzano, G., Scaglia, B., Adani, F. (2009). Substituting energy crops with organic fraction of municipal solid waste for biogas production at farm level: A full-scale plant study. Process Biochemistry, doi:10.1016/j.procbio.2009.03.014.

Rico, J.L., Garcia, H., Rico, C. and Tejero, I. (2007). Characterisation of solid and liquid fractions of dairy manure with regard to their component distribution and methane production.

Rubæk, G.H., Henriksen, K., Petersen, J., Rasmussen, B., Sommer, S.G. (1996). Effects of application technique and anaerobic digestion on gaseous nitrogen loss from animal slurry applied to ryegrass (Lolium perenne). Journal of Agricultural Science 126, 481–492.

Preusch, P.L., Adler, P.R., Sikora, L.J., Tworkoski, T.J. (2002). Nitrogen and phosphorus availability in composted and uncomposted poultry litter. Journal of Environmental Quality 31, 2051–2057.

Raviv, M., Medina, S., Krasnovsky, A., Ziada, H. (2004). Organic matter and nitrogen conservation in manure compost for organic agriculture. Compost Science and Utilisation 12, 6–10.

Rao Bhamidimarri, S.M., Pandey, S.P. (1996). Aerobic thermophilic composting of piggery solid wastes. Water Science and Technology 33 (8), 89-94.

Rotenberg, D., Cooperband, L., Stone, A. (2005). Dynamic relationship between soil properties and foliar disease as effected by annual additions of organic amendment to a sandy soil vegetable production systems. Soil Biology and Biochemistry 37, 1343–1357.

Rovira, P., Vallejo, V.R. (2002). Labile and recalcitrant pools of carbon and nitrogen in organic matter decomposing at different depths in soil: an acid hydrolysis approach. Geoderma 107, 109–141.

Rubæk GH, Henriksen K, Petersen J, Rasmussen B, Sommer SG. 1996. Effects of application technique and anaerobic digestion on gaseous nitrogen loss from animal slurry applied to ryegrass (Lolium perenne). Journal of Agricultural Science 126, 481–492.

Russell (1988). Chapter 20 Plant nutrients in soil: nitrogen. In “Russell‟s soil conditions & Plant growth”. Ed. Wild A. Pub. Longman.


124

Sánchez, M.M., Gomez, X., Barriocanal, G., Cuetos, M.J. and Moran, A. (2008) assessment of the stability of livestock farm wastes treated by anaerobic digestion. Int. Biodeterioration and Biodegradation, 62, p421.

Sánchez-Monedero, M.A., Mondini, C., de Nobili M, Leita, L., Roig, A. (2004). Land application of biosolids. Soil response to different stabilization degree of the treated organic matter. Waste Management 24, 325–332.

Schröder, J.J., Jansen, A.G., Hilhorst, G.J. (2005). Long-term nitrogen supply from cattle slurry. Soil Use and Management 21, 196–204.

Shao, Z-H., He, P-J., Zhang, D-Q. and Shao, L-M. (2009). J Haz. Materials, 164, p1191.

Sharpley, A., Moyer, B. (2000). Phosphorus forms in manure and compost and their release during simulated rainfall. J. Environ. Qual. 29, 2053.

SI (1989). Stutory Instrument No. 1263. The Sludge (Use in Agriculture) Regulations 1989. http://opsi.gov.uk/si/si1989/uksi_19891263_en_1.htm.

Sikora, L.J., Enkiri, N.K. (2000). Efficiency of compost-fertilizer blends compared with fertilizer alone. Soil Science 165 (5), 444–451.

Six, J., Bossuyt, H., Degryze, S., Denef, K. (2004). A history of research on the link between (micro)aggregates, soil biota, and soil organic matter dynamics. Soil and Tillage Research 79, 7–31.

Skiba, U., Hargreaves, K.J., Fowler, D., Smith, K.A. (1992). Fluxes of nitric and nitrous oxides from agricultural soils in a cool temperate climate. Atmospheric Environment 26, 2477-2488.

Smith, P., Powlson, D.S., Smith, J.U., Falloon, P., Coleman, K. (2000). Meeting Europe‟s climate change commitments: quantitative estimates of the potential for carbon mitigation by agriculture. Global Change Biology 6, 525–539.

Smith (2008) The Implications for Human Health and the Environment of Recycling Biosolids on Agricultural Land

Sommer, S.G., Genermont, S., Cellier, P., Hutchings, N.J., Olesen, J.E., Morvan, T. (2003). Processes controlling ammonia emission from livestock slurry in the field. European Journal of Agronomy 19, 465–486.

Sommer, S.G., Husted, S. (1995). The chemical buffer system in raw and digested animal slurry. Journal of Agricultural Science, Cambridge 124, 45-53.

Tam, N.F.Y., Tiquia, S.M. (1999). Nitrogen transformation during co-composting of spent pig manure, sawdust litter and sludge under forced-aerated system. Environmental Technology 20, 259-267.

Tejada, M., Hernandez, M.T. Garcia C. (2009). Soil restoration using composted plant residues: Effects on soil properties. Soil & Tillage Research 102, 109–117

Tiquia, S.M., Tam, N.F.Y. (2000). Fate of nitrogen during composting of chicken litter. Environmental Pollution 110, 535-541.


125

Tognetti, C., Mazzarino, M.J., Laos, F. (2008). Compost of municipal organic waste: Effects of different management practices on degradability and nutrient release capacity. Soil Biology & Biochemistry 40 2290–2296.

Trinsoutrot, I., Nicolardot, B., Justes, E.S. (2000). Decomposition in the field of residues of oilseed rape grown at two levels of N fertilisation: effects on the dynamics of soil mineral N between successive crops. Nutrient Cycling in Agroecosystems 56, 125–137.

Trzcinski, A.P. and Stuckey, D.C. (2009). Continuous treatment of the organic fraction of municipal solid waste in an anaerobic two-stage membrane process with liquid recycle. Water Res., 43, p2449.

UKWIR (2006). Application of phosphorous in industrial biosolids to agricultural soils. Final report for UKWIR, Defra and EA on UKWIR project SL02 B.

Vaca-Paulin, R., Esteller-Alberich, M.V., Lugo-de la Fuente, J. and Zavaleta-Mancera, H.A. (2006). Effect of sewage sludge or compost on the sorption and distribution of copper and cadmium in soil. Waste Management, 26, p71-81.

Vallejo, A., Skiba, U.M., Garcia-Torres, L., Arce, A., Lopez-Fernandez, S., Sanchez-Martin, L. (2006). Nitrogen oxides emission from soils bearing a potato crop as influenced by fertilization with treated pig slurries and composts. Soil Biology and Biochemistry 38, 2782–2793.

Venglovsky, J., Saakova, N. and Placha, I. (2009). Pathogens and antibiotic residues in animal manures and hygienic and ecological risks related to subsequent land application. Bioresource Technology, in press.

Wagland, S.T., Tyrrel, S.F., Godley, A.R. and Smith, R. (2009). Test methods to aid in the evaluation of the diversion of biodegradable municipal waste (BMW) from landfill. Waste Management, 29, 1218-1226.

Webb, J., Henderson, D., Anthony, S.G. (2001). Optimising livestock manure applications to reduce nitrate and ammonia pollution : scenario analysis using the MANNER model. Soil Use and Management. 17, 188-94.

Webb, J., Misselbrook, T.H. (2004) A mass-flow model of ammonia emissions from UK livestock production. Atmospheric Environment 38, 2163-2176.

Weber, J., Karczewska, A., Drozd, J., Licznar, M., Licznar, S., Jamroz, E., Kocowicz, A. (2007). Agricultural and ecological aspects of a sandy soil as affected by the application of municipal solid waste composts. Soil Biol. Biochem., 39, p1294.

WESA and Bates. (2002). Guidelines for the Utilization of Pulp and Paper Mill Biosolids on Agricultural Lands. Ontario Forest Industries Association. Toronto: Ontario Withers, P.J.A., Edwards, A.C. and Foy, R.H. (2001). Phosphorus cycling in UK agriculture and implications for phosphorous loss from soil. Soil Use and Management. 17, p139-149.

Witter, E., Lopez-Real, J. (1988). Nitrogen losses during the composting of sewage sludges, and the effectiveness of clay soil, zeolite, and compost in adsorbing the volatilized ammonia. Biological Wastes 23, 279-294.


126

Wolkowski, R.P. (2003). Nitrogen management considerations for landspreading municipal solid waste compost. Journal of Environmental Quality 32, 1844–1850.

WRc (2003) Pathogens in biosolids – Microbiological Risk Assessment

Wulf, S., Maeting, M., Clemens, J. (2002). Application technique and slurry co-fermentation effects on ammonia, nitrous oxide, and methane emissions after spreading: I. Ammonia volatilisation. Journal of Environmental Quality 31, 1789-1794.

Wulf, S., Maeting, M., Clemens, J. 2002b. Application technique and slurry co-fermentation effects on ammonia, nitrous oxide, and methane emissions after spreading: II. Greenhouse gas emissions. Journal of Environmental Quality 31, 1795–1801.

Yamulki, S., Goulding, K.W.T., Webster, C.P., Harrison, R.M. (1995). Studies on NO and N2O fluxes from a wheat field. Atmospheric Environment 29, 1627-1635.

Zhang, R., El-Mashad, H.M., Hartman, K., Wang, F., Liu, G., Choate, C. and Gamble, P. Characterization of food waste as feedstock for anaerobic digestion. Bioresource Technol., 98, p929.

NRAES (1992). On Farm Composting Handbook (NRAES-54), NRAES (Northeast Regional Agricultural Engineering Service): Ithica, NY. <http://compost.css.cornell.edu/OnFarmHandbook/coverpg.html


127

A11. APPENDIX

A11.1 Characteristics of waste types used in this assessment (mean values)

Dry Organic matter Total RAN Organic Calc Total Total Total Total Metals all mg/kg DM

Waste Treatment Code matter pH LOI TOC (cal) N NH4+ (NH3+NO3) N PAN P K Mg S Cd Cr Cu Pb Hg Ni Zn

% w et w t g/kg DM g/kg dm g/kg DM g/N kg DM g N/kg DM g N/kg DM g N/kg DM g P/kg DM g K/kg DM g /kg DM g /kg DM

Cattle Slurry 11 23.7 7.8 700 406 406 36.7 23.57 23.57 13.1 26.7 43.80 41.5 2.0 0.5 6.3 76 8 7 215

Pig Slurry 12 12.9 7.5 645 352 74.0 49.02 49.02 25.0 59.0 32.43 44.9 12.3 0.4 3.2 589 4 18 1558

Cattle manure 13 16.1 6.1 450 450 27.9 3.90 24.0 11.6 10.35 19.1 4.7 0.3 7.4 57 4 0.070 4 245

Pig manure 14 20.7 6.3 367 367 46.0 46.0 18.4 7.30 16.6 1.3 0.4 2.0 422 3 11 446

Sheep manure 15 53.3 6.5 418 418 46.0 2.50 43.5 19.9 3.50 10.0

Poultry manure 16 40.0 7.7 552 552 34.5 3.50 31.0 13.4 24.40 11.9 0.6 15.4 70 15 10 359

Horse Manure 17 51.5 8.0 427 427 27.0 3.50 23.5 11.0

Livestock slurry 18 24.1 6.6 795 435 29.3 10.01 10.01 19.3 14.6 20.68

Manures (general) 19 38.7 9.0 700 383 16.4 16.4 6.6 4.50 3.4 1.8 0.2 15.0 41 13 0.070 19 213

Sludges from biological treatment of industrial waste 23 11.9 7.6 500 273 4.7 0.83 0.83 3.9 0.8 0.41 2.5 2.6 0.8 22.3 12 7 3.937 15 14

Sewage sludge (general) 26 9.5 7.0 589 334 334 38.7 7.88 9.01 29.7 23.3 21.69 5.4 2.8 2.2 82.1 432 145 2.167 38 675

Plant tissue waste 31 47.1 7.7 393 221 221 13.3 0.29 0.53 12.8 4.6 2.78 12.8 2.0 0.5 10.9 30 52 0.228 8 121

Garden and park waste 32 51.4 8.4 361 260 260 24.1 0.44 0.49 23.6 9.9 2.19 11.6 2.8 0.5 22.7 53 88 0.523 13 167

Seperately collected fraction MSW (curb side collections) 33 49.2 7.2 575 300 300 18.9 0.39 0.40 18.5 7.8 2.90 11.7 2.9 0.6 21.1 52 98 0.241 12 160

CA Site greenwaste 34 47.0 7.1 737 403 13.3 0.00 0.00 13.3 3.2 0.00

Grass cuttings 35 36.9 8.3 864 472 23.5 0.84 0.84 22.7 9.9 3.63 12.8 1.4 0.1 6.2 35 17 0.120 6 124

Source segregated green waste 36 60.1 6.5 393 220 220 12.3 0.10 0.14 12.1 4.0 2.21 7.9 3.5 0.6 22.4 53 106 0.256 16 188

Wood 37 46.5 7.9 880 481 481 20.1 20.1 2.4 0.40 1.5 2.0 1.3 64.2 148 2016 0.361 16 858

Greenwaste (general) 38 59.4 6.0 438 268 268 17.6 0.70 0.82 16.7 6.2 2.23 7.9 2.5 0.6 23.0 61 134 0.288 14 192

Dairy Production Waste 40.1 2.7 6.1 459 251 67.1 13.30 13.59 53.5 51.1 10.81 12.1 2.4 2.1 21.3 86 15 9.239 26 145

ABPR waste 40.4 28.7 6.2 700 383 20.9 2.87 2.87 18.0 2.9 3.83 10.8 1.0 0.4 17.3 42 67 0.145 30 170

Foodwaste (general) 40.5 7.5 6.0 513 280 38.9 17.10 18.02 20.9 29.7 9.62 10.6 2.8 14.9 88.5 223 35 3.477 79 251

Sludges from washing and cleaning for food preperation 41 12.5 7.5 538 294 56.4 29.24 30.31 26.1 42.8 7.28 11.9 2.4 12.3 114.8 221 58 1.899 77 631

Gut content 42 18.8 4.6 740 404 20.5 2.75 2.82 17.7 8.1 12.22 17.9 2.4 0.1 1.5 10 2 0.046 2 66

Vegetable washings 44 1.4 5.2 553 302 41.5 7.41 7.55 33.9 27.9 0.00 60.2 3.6

Vegetable production waste (peelings, choppings) 45 8.3 4.4 893 488 24.2 1.56 1.56 22.6 9.7 3.99 38.3 1.6

Sugar processing 46 46.4 5.6 317 173 9.7 0.29 0.29 9.4 4.8 1.96 1.5 2.8 2.9 33.9 59 15 0.036 31 37

Wastes from baking and confectionairy 47 9.5 5.0 530 290 16.3 4.18 4.20 12.1 7.8 4.77 5.8 1.6 3.3 67.2 123 47 0.583 74 486

Beverage production 48 4.2 7.2 509 278 32.9 2.50 2.53 30.3 17.7 5.19 8.3 3.3 34.6 49.2 86 46 7.167 46 296

Fish farm waste (faeces and uneaten food) 49 38.9 8.6 164 89 9.6 0.21 0.23 9.4 5.9 4.81 1.4 2.0

Green waste and COM 50.6 58.0 7.6 527 306 306 0.09 0.10 -0.1 0.1 2.36 10.3 0.8 19.7 69 107 0.198 13 176

MSW and manure 50.7 11.7 7.2 500 273 72.1 37.27 37.27 34.8 54.7

Bio waste municipal 51 36.1 8.6 627 323 323 42.0 21.58 21.73 20.3 28.2 5.49 14.3 3.6 0.5 29.7 84 86 0.142 75 232

Biodegradable kitchen and canteen wastes 52 49.3 7.9 530 308 308 0.65 0.65 -0.6 0.6 25.0 2 12

BMW 53 21.2 7.4 504 275 47.1 31.95 32.00 15.1 38.0 4.59 9.2 6.1 1.3 254.8 421 407 0.586 121 791

MBT residues 54 58.9 505 207 207 9.2 0.51 0.52 8.7 1.8 0.79 3.3 1.6 2.0 78.3 158 315 0.746 56 472

Manure and biobin material 55 35.0 7.7 700 383 51.9 3.00 48.9 22.6 6.54 19.1 6.4 0.5 19.3 133 23 0.140 20 436

De-inking sludges from paper recycling 71 39.3 7.5 399 227 227 4.4 0.68 0.68 3.7 0.7 1.06 1.2 2.2 0.3 12.8 88 8 0.143 6 66

Dredgings 73 37.6 7.8 98 54 6.5 0.18 0.19 6.3 0.8 0.97 2.4 2.8 0.7 26.0 72 190 0.414 25 210

Construction and demolition wastes (soil) 74 46.2 54 30 0.4 0.4 0.2 0.31 8.0 16.2 0.8 27.2 26 39 0.537 26 68

Chipboard 75 95.0 419 419 32.7 32.7 16.4 0.2 3.8 8 17 0.010 2 22

MDF 76 84.0 580 580 56.0 56.0 11.2 0.07 0.5 0.1 0.8 3 1 0.060 5

Sludges from treatment of drinking water 78 5.7 50 27 33.8 0.35 0.35 33.4 16.9 17.50 1.5 33.8 432 73 1.570 28 543

Not specified N 41.1 404 390 390 40.6 8.42 8.58 32.0 21.4 2.96 12.8 1.6 0.4 18.1 57 66 0.154 12 176


128

A11.2 Characteristics of waste treatments used in this assessment (mean values)

Dry Organic matter Total RAN Organic Calc Total Total Total Total Metals all mg/kg DM

Waste Treatment Code matter pH LOI TOC (cal) N NH4+ (NH3+NO3) N PAN P K Mg S Cd Cr Cu Pb Hg Ni Zn

% w et w t g/kg DM g/kg dm g/kg DM g/N kg DMg N/kg DM g N/kg DM g N/kg DM g P/kg DM g K/kg DM g /kg DM g /kg DM

Composted and Lime CL 49.6 9.5 271 148 23.2 23.2 2.3 2.8 10.8 3.1 3.2 0.5 17 51 106 0.18 14 178

Composted mechanical dewatered CM 50.4 8.9 325 178 18.4 18.4 1.8 2.9 11.5 2.9 3.6 0.6 20 59 75 0.15 12 164

Compost C 54.4 8.0 232 127 20.2 0.22 0.22 20.0 2.2 3.6 8.5 2.8 7.8 0.6 25 79 142 0.35 18 221

Meso AD and mechanical dewater DM 22.7 43.3 43.3 24.2 4.3 135 623 308 3.01 60 939

Mesophilic AD D 5.0 7.8 52.7 52.7 29.6 1.5 2.5 10.4 3.0 194 547 222 2.75 55 878

Heat dried H 91.4 5.9 396 216 52.3 52.3 31.4 24.8 19.9 3.8 84 552 873 1.55 40 1088

Liquid storage LS 8.9 231 126 38.2 38.2 22.9 19.3 1.8 93 339 164 1.58 41 717

Lime stabilised L 6.7 12.5 41.3 41.3 16.5 1.6 179 413 116 2.01 26 667

Mechanical dewatered M 23.9 34.8 0.73 34.0 21.1 17.0 2.1 1.9 83 461 204 2.14 43 693

Not specified N 19.6 6.5 263 144 28.3 28.3 6.6 6.4 2.5 12.9 6.2 44 148 44 2.62 50 267

Liquid fraction Q 2.1 3.2 50 420 137 31 527

AD + Lime stabilisation RL 8.0 69.4 69.4 8.7 22.9 6.4 1.0 51 196 173 0.47 35 443

AD R 15.0 7.6 181 99 66.0 12.88 53.1 44.8 6.8 19.7 3.7 6.5 10.2 114 255 110 1.27 65 437

Thermo AD and mechanical dewat TM 19.1 51.2 51.2 14.4 2.6 170 354 1003 2.13 28 901

Thermophilic AD T 12.5 3.9 36.7 36.7 26.3 5.1 0.4 2.0 1.7 37 409 194 4.69 24 658

Untreated U 5.6 5.1 279 152 36.0 36.0 14.4 20.4 29.7 3.3 6.0 1.7 50 405 107 2.04 32 588


129

A11.3 Heavy metal loadings from waste types applied to give a PAN of 100 kg N/ha

Cd Cr Cu Pb Hg Ni Zn

Number Application Number Application Number Application Number Application Number Application Number Application Number Application

Application kg/ha.yr Application kg/ha.yr Application kg/ha.yr Application kg/ha.yr Application kg/ha.yr Application kg/ha.yr Application kg/ha.yr

Waste Code to soil limit 0.15 to soil limit 3 to soil limit 7.5 to soil limit 15 to soil limit 0.1 to soil limit 3 to soil limit 15

Sewage sludge (general) 26 591 0.011 2539 0.42 151 2.23 907 0.7 241 0.011 772 0.19 89 3.5

Cattle Slurry 11 3855 0.002 45512 0.02 1177 0.29 23875 0.0 5587 0.03 384 0.8

Pig Slurry 12 10115 0.001 200944 0.01 337 1.00 101925 0.0 4945 0.03 117 2.6

Cattle manure 13 2887 0.002 16981 0.06 685 0.49 18563 0.0 4471 0.001 3982 0.04 146 2.1

Pig manure 14 3327 0.002 98469 0.01 147 2.29 43242 0.0 2610 0.06 127 2.4

Sheep manure 15

Poultry manure 16 1435 0.005 9369 0.11 643 0.52 5981 0.1 1962 0.08 115 2.7

Horse Manure 17

Livestock slurry 18

Manures (general) 19 2165 0.003 4710 0.23 538 0.63 3421 0.2 2530 0.001 518 0.29 95 3.2

Sludges from biological treatment of industrial waste 23 69 0.096 403 2.67 235 1.43 767 0.9 6 0.473 82 1.84 186 1.7


Plant tissue waste 31 664 0.010 4541 0.24 515 0.65 607 1.1 545 0.005 901 0.17 118 2.6

Garden and park waste 32 1233 0.005 4711 0.23 634 0.53 767 0.9 514 0.005 1106 0.14 184 1.7

Seperately collected fraction MSW (kerb side collections) 33 863 0.008 3983 0.27 507 0.66 540 1.3 873 0.003 966 0.16 150 2.1

CA Site greenwaste 34

Grass cuttings 35 6535 0.001 17313 0.06 951 0.35 3949 0.2 2228 0.001 2701 0.06 247 1.2

Source segregated green waste 36 429 0.015 1932 0.56 254 1.32 257 2.6 424 0.006 377 0.40 66 4.7

Wood 37 121 0.055 405 2.66 55 6.11 8 83.5 181 0.015 222 0.68 9 35.6

Greenwaste (general) 38 659 0.010 2897 0.37 339 0.99 313 2.2 580 0.005 659 0.23 99 3.1

Dairy Production Waste 40.1 1607 0.004 25828 0.04 1987 0.17 23055 0.0 149 0.018 2945 0.05 1092 0.3

ABPR waste 40.4 426 0.016 1792 0.60 231 1.45 290 2.3 534 0.005 145 1.03 52 5.9

Foodwaste (general) 40.5 132 0.050 3619 0.30 448 0.75 5733 0.1 231 0.012 561 0.27 365 0.8

Sludges from washing and cleaning for food preperation 41 229 0.029 4020 0.27 651 0.52 4998 0.1 609 0.004 837 0.18 210 1.5

Gut content 42 4929 0.001 59646 0.02 2780 0.12 27940 0.0 4819 0.001 7542 0.02 380 0.8

Vegetable washings 44

Vegetable production waste (peelings, choppings) 45

Sugar processing 46 110 0.060 1527 0.71 276 1.22 2140 0.3 3636 0.001 230 0.65 401 0.8

Wastes from baking and confectionairy 47 157 0.042 1253 0.86 214 1.57 1121 0.6 362 0.007 159 0.94 50 6.2

Beverage production 48 34 0.195 3873 0.28 695 0.48 2636 0.3 67 0.041 573 0.26 185 1.7

Fish farm waste (faeces and uneaten food) 49

Green waste and COM 50.6

MSW and manure 50.7

Bio waste municipal 51 3477 0.002 10230 0.11 1127 0.30 2214 0.3 5356 0.001 562 0.27 376 0.8

Biodegradable kitchen and canteen wastes 52

BMW 53 1892 0.003 1608 0.67 303 1.11 634 1.1 1753 0.002 470 0.32 149 2.1

MBT residues 54 62 0.107 252 4.28 39 8.66 39 17.2 66 0.041 49 3.06 12 25.8

Manure and biobin material 55 3104 0.002 12575 0.09 572 0.59 6722 0.1 4353 0.001 1725 0.09 160 1.9

De-inking sludges from paper recycling 71 147 0.045 572 1.88 26 12.94 582 1.2 128 0.021 182 0.82 32 9.7

Dredgings 73 74 0.090 342 3.15 38 8.73 30 23.0 54 0.050 49 3.08 12 25.5

Construction and demolition wastes (soil) 74 17 0.396 81 13.26 27 12.64 35 19.2 10 0.262 12 12.63 9 33.3

Chipboard 75 5995 0.001 46832 0.02 6807 0.05 6445 0.1 44145 0.000 12263 0.01 2350 0.1

MDF 76 6204 0.001 144627 0.01 14356 0.02 106221 0.0 5076 0.001 70501 0.00 7084 0.0

Sludges from treatment of drinking water 78 754 0.009 5385 0.20 131 2.56 1563 0.4 291 0.009 899 0.17 96 3.2

Not specified N 3576 0.002 12733 0.08 1252 0.27 2201 0.3 3747 0.001 2685 0.06 376 0.8


130

A11.4 Heavy metal loadings from waste treatments applied to give a PAN of 100 kg N/ha


Number ApplicationNumber Application Number Application Number Application Number Application Number Application Number Application



sewage sludge (general) 26 591 0.011 2539 0.42 151 2.2 907 0.7 241 0.011 772 0.19 89 3.5

Composted and Lime CL 291 0.023 1476 0.73 152 2.2 148 4.6 357 0.008 245 0.61 40 7.7

Composted mechanical dewateredCM 202 0.033 996 1.08 105 3.2 166 4.1 331 0.008 239 0.63 35 8.9

Compost C 227 0.029 959 1.12 94 3.6 106 6.4 171 0.016 185 0.81 31 10.0

Meso AD and mechanical dewaterDM

Mesophilic AD D

Heat dried H 543 0.012 4023 0.27 191 1.8 244 2.8 547 0.005 1177 0.13 89 3.5

Liquid storage LS 862 0.008 2658 0.41 227 1.5 948 0.7 390 0.007 842 0.18 99 3.1

Lime stabilised L

Mechanical dewatered M 744 0.009 2735 0.39 154 2.2 701 1.0 267 0.010 731 0.21 94 3.3

Not specified N

Liquid fraction Q

AD + Lime stabilisation RL

AD R 291 0.023 4232 0.25 591 0.6 2770 0.2 953 0.003 1035 0.14 317 1.0

Thermo AD and mechanical dewatTM

Thermophilic AD T

Untreated U 553 0.012 3086 0.35 119 2.8 913 0.7 191 0.014 685 0.22 76 4.1


131

A11.5 Heavy metals loadings from waste types applied as soil conditioner


Number Application Number Application Number Application Number Application Number Application Number Application Number Application




Cattle Slurry 11 117 0.056 1383 0.78 36 9.39 725 0.9 170 0.88 12 26.5

Pig Slurry 12 121 0.055 2401 0.45 4 83.51 1218 0.6 59 2.54 1 221.1

Cattle manure 13 224 0.029 1319 0.82 53 6.32 1441 0.5 347 0.008 309 0.49 11 27.2

Pig manure 14 133 0.050 3928 0.27 6 57.49 1725 0.4 104 1.44 5 60.8

Sheep manure 15

Poultry manure 16 118 0.056 772 1.40 53 6.35 493 1.4 162 0.93 9 32.5

Horse Manure 17

Livestock slurry 18

Manures (general) 19 252 0.026 549 1.96 63 5.36 399 1.7 295 0.009 60 2.48 11 27.8

Sludges from biological treatment of industrial waste 23 38 0.175 220 4.89 128 2.62 419 1.6 3 0.865 45 3.36 102 3.0


Plant tissue waste 31 64 0.104 436 2.47 49 6.80 58 11.6 52 0.052 86 1.74 11 27.4

Garden and park waste 32 64 0.102 246 4.37 33 10.13 40 16.9 27 0.101 58 2.59 10 32.1

Seperately collected fraction MSW (curb side collections) 33 55 0.119 256 4.21 33 10.34 35 19.6 56 0.048 62 2.42 10 32.1

CA Site greenwaste 34

Grass cuttings 35 519 0.013 1376 0.78 76 4.45 314 2.2 177 0.015 215 0.70 20 15.7

Source segregated green waste 36 47 0.141 211 5.10 28 12.09 28 24.1 46 0.058 41 3.63 7 42.8

Wood 37 64 0.103 215 5.01 29 11.51 4 157.3 96 0.028 118 1.27 5 66.9

Greenwaste (general) 38 57 0.116 251 4.29 29 11.43 27 25.0 50 0.054 57 2.63 9 35.9

Dairy Production Waste 40.1 8 0.836 127 8.49 10 34.42 113 6.0 1 3.683 14 10.37 5 57.6

ABPR waste 40.4 57 0.116 239 4.51 31 10.91 39 17.5 71 0.038 19 7.74 7 44.5

Foodwaste (general) 40.5 1 5.306 34 31.54 4 79.55 54 12.5 2 1.240 5 28.33 3 89.6

Sludges from washing and cleaning for food preperation 41 2 3.144 37 29.25 6 56.39 46 14.8 6 0.484 8 19.58 2 160.7

Gut content 42 409 0.016 4945 0.22 230 1.46 2317 0.3 400 0.007 625 0.24 31 9.8

Vegetable washings 44

Vegetable production waste (peelings, choppings) 45

Sugar processing 46 5 1.249 73 14.69 13 25.37 103 6.6 175 0.015 11 13.59 19 16.0

Wastes from baking and confectionairy 47 10 0.679 77 13.92 13 25.38 69 9.8 22 0.121 10 15.29 3 100.6

Beverage production 48 1 7.454 102 10.60 18 18.43 69 9.8 2 1.545 15 9.99 5 63.7

Fish farm waste (faeces and uneaten food) 49

Green waste and COM 50.6 48 0.138 335 3.22 30 11.28 39 17.5 84 0.032 73 2.04 11 28.7

MSW and manure 50.7

Bio waste municipal 51 79 0.083 234 4.61 26 13.04 51 13.4 122 0.022 13 11.67 9 36.0

Biodegradable kitchen and canteen wastes 52 464 0.72 80 3.9

BMW 53 23 0.289 19 55.51 4 91.79 8 88.6 21 0.128 6 26.45 2 172.3

MBT residues 54 17 0.378 71 15.16 11 30.70 11 61.0 19 0.144 14 10.86 3 91.4

Manure and biobin material 55 105 0.063 426 2.53 19 17.33 228 3.0 148 0.018 58 2.57 5 56.9

De-inking sludges from paper recycling 71 98 0.067 383 2.81 17 19.33 390 1.7 85 0.032 122 1.23 21 14.5

Dredgings 73 14 0.459 67 16.13 8 44.70 6 117.6 11 0.257 9 15.80 2 130.4

Construction and demolition wastes (soil) 74 6 1.032 31 34.54 10 32.94 14 50.0 4 0.682 5 32.91 4 86.7

Chipboard 75 410 0.016 3200 0.34 465 0.72 440 1.5 3017 0.001 838 0.18 161 1.9

MDF 76 857 0.008 19989 0.05 1984 0.17 14681 0.0 702 0.004 9744 0.02 979 0.3

Sludges from treatment of drinking water 78 2 2.708 17 61.85 0 790.56 5 134.1 1 2.873 3 51.61 0 993.7

Not specified N 109 0.061 387 2.78 38 8.83 67 10.1 114 0.024 82 1.84 11 27.0


132

A11.6 Heavy metal loadings from waste treatments applied as soil conditioner


Number ApplicationNumber Application Number Application Number Application Number Application Number Application Number Application



sewage sludge (general) 26 14 0.486 58 18.46 3 97.2 21 32.5 6 0.487 18 8.46 2 151.8

Composted and Lime CL 56 0.118 283 3.81 29 11.5 28 23.9 68 0.039 47 3.19 8 40.1

Composted mechanical dewateredCM 59 0.113 288 3.74 30 11.1 48 14.1 96 0.028 69 2.17 10 30.8

Compost C 39 0.170 164 6.55 16 20.9 18 37.3 29 0.092 32 4.72 5 58.3

Meso AD and mechanical dewaterDM

Mesophilic AD D

Heat dried H 5 1.323 37 29.12 2 191.3 2 302.4 5 0.537 11 13.87 1 377.0

Liquid storage LS 6 1.042 20 55.13 2 201.3 7 97.3 3 0.941 6 24.25 1 426.2

Lime stabilised L

Mechanical dewatered M

Not specified N 2 3.235 47 22.88 4 77.5 30 22.8 2 1.370 6 26.30 2 139.4

Liquid fraction Q

AD + Lime stabilisation RL

AD R 1 7.704 12 86.39 2 193.0 8 83.1 3 0.962 3 49.19 1 331.3

Thermo AD and mechanical dewatTM

Thermophilic AD T

Untreated U 10 0.677 54 19.80 2 159.6 16 42.1 3 0.803 12 12.43 1 232.0

Documents

DEFRA WQ0206: AGRONOMIC BENEFITS AND …sciencesearch.defra.gov.uk/Document.aspx?Document=DEFRA8021App… · sample that can be combusted at 550oC. The loss in weight during combustion