Chapter 9 Fate of contaminants in soil

245

Chapter 9

FATE OF C O N T A M I N A N T S IN SOIL W. Peijnenburg

Abstract This chapter discusses the present status of our knowledge on the nature and behaviour of contaminants in the soil. The fate of contaminants is described according to the processes involved in natural attenuation: biodegradation, diffusion, dilution, sorption, volatilization and chemical and biochemical stabilization. The role of the uptake by vegetation and animals in attenuation is also considered. The contaminants are arranged in the structural groups heavy metals, nutrients, and polyaramites such as the chlorinated ones. Since the leading theory on the uptake of chemicals by soil organisms is the equilibrium-partitioning (EP) theory, its theoretical aspects and practical applications and limitations are discussed in detail. This provides us with the means to apply this concept to the relevant bioavailability and bioaccessibility fractions of the total contamination content in soil. The defined liquid, particulate and biotic phase play a leading role in the assessment of this EP. Three different tables contain overviews of commonly used extraction techniques for the three different contaminant groups. They contain extraction methods, endpoints of consideration and the soil properties of interest as organic matter content, pH, redox potential, clay content and dissolved organic carbon. Many controlling factors can be quantified and modelled on a routine basis. Unfortunately the various biotic species seem to differ in their uptake procedures. Therefore, there is no single expression for the bioavailable or bioaccessible fractions of soil contamination. It is concluded that monitoring the effective presence of contaminants should be based on specific extraction techniques, closely related to biological availability.

9.1 Introduction M a n y varieties of organic a n d inorganic c o m p o u n d s are present in soils. Due to natural

and

anthropogenic

emission,

point sources

and

veils of various

chemicals are present in soil on mondial, continental, regional a n d local scales. We will discuss the fate of these substances in the soil in the context of the complexity of soil a n d the large variety of substances i n t r o d u c e d into it. These substances or c o m p o n e n t s are defined as c o n t a m i n a n t s w h e n they are

W.J.G.M. Peijnenburg

246

present in elevated concentrations. For reasons of simplicity and clarity, they have been divided into three structural groups: nutrients, (persistent) organic chemicals (including pesticides) and heavy metals. After an introduction to the soil environment, contaminants and their properties, the interaction between the contaminants and the soil matrix and its consequences will be described according to the definition of natural attenuation. The aspects of bioavailability and bioaccessibility, which may be seen as the key elements governing the persistence and the potential and actual exposure of biota to contaminants will be discussed. A monitoring strategy for bioavailable and bioaccessible contaminants present in soils is described in the final sections. This includes a brief summary of the methods available for measuring chemicals in the field. 9.2 Soil Environment

Soil represents one of the most complex matrices in the environment because of its heterogeneity. It is an interplay of particles of different size, organic matter of different quality and quantity, pore water, (pore) air, and biota. It differs radically with regard to size as well as ecophysiological characteristics and it may lead to conditions that vary from aerobic to anaerobic. This interplay has been depicted schematically in Figure 1. Clay- organic matter complex

Condensed Open pore

Bacteria Quartz

:.%. Pore opening

Closed pore

Clay domain Organic fnatter sesqui oxides

Fungal hyphae

Figure 1. Microscopic associations between organisms, organic matter and mineral particles (modified after Paul and Clark, 1989, and Cuypers, 2001)

Fate of contaminants in soil

247

Following interactions at micro- and macro-scales, soils act as sinks for a blend of organic and inorganic compounds of natural and anthropogenic origin. As such the soil provides large quantities of essential and non-essential elements, and compounds of a widely varying nature as sources of nutrition to biota. Contaminants are introduced into the soil by natural as well as anthropogenic sources. The contaminant groups discussed are persistent organics as pesticides (POPs) and hydrocarbons, heavy metals and nutrients. Since the current international focus is on POPs, they will receive the most attention here. 9.2.1 Polyaromatic hydrocarbons (PAHs)

Volcanic activities, hydrothermal wells and forest fires in combination with the burning of fossil fuels are examples of important natural sources of PAHs due to incomplete combustion of organic material. In soil, background concentrations in pristine samples can range from 50 to 1100 ~tg.kg-1 (Fritz and Engst, 1975). Concentrations of PAHs in contaminated soils, due to wood preservation, creosote production, gas works etc, may range upward to some thousands mg.kg -1 (Wilson and Jones, 1993).The concentrations in the subsoils differ from the top soils. Here we focus on top soils. 9.2.2 Persistent organic pollutants (POPs)

Another group of organic contaminants which are present in many soils are polychlorinated biphenyls (PCBs). PCBs were first described in 1881 but have been produced technically to be used as additives in plastics, in electrical insulators, as basis compounds for pesticides, etc. Because of their varied use they have been spread to all environmental compartments and have been introduced into soils through various routes including application of pesticides, via refuse or via deposition of air particles. Although PCB production has been reduced in most countries since the 1970s and production was officially stopped in 1993, the total PCB content in background soil today amounts to 21,000 tons. A group of chemicals classified as being among the most persistent and most toxic compounds ever produced by mankind, are the chlorinated dibenzo-pdioxins and dibenzofurans. These chemicals are produced by biomass burning, but they are also the by-products of the large-scale production of chlorinated aromatic compounds (Alcock et al., 1998). Kjeller et al. (1991) measured the concentrations of chlorinated dibenzo-p-dioxins and dibenzofurans in soil samples from the United Kingdom taken from 1850 to the present day (Figure 2). Concentrations of 30-60 ng.kg -1 before 1945 are related to the large-scale combustion of fuels (coal) and the smelting of metals for the production of iron

W.J.G.M. Peijnenburg

248

140 - -

[]

120 - -

[] []

100 - -

[]

r

!

[]

80--

9

60 m

o

r

[]

[]

40--

20--

0 1820

I

I

I

I

I

I

I

I

[

1840

1860

1880

1900

1920

1940

1960

1980

2000

Year

Figure 2. The concentrations of chlorinated dibenzo-p-dioxins and dibenzofurans in

soil samples taken from the UK since 1850 (modified after Kjeller et al., 1991)

for an agriculture quantities. Examples of this include the group of phthalate esters, a group of chemicals that is widely used in enormous quantities as plasticizers. Most phthalate esters are difficult to quantify at levels exceeding the detection limits of the currently available analytical techniques. Also, nonhydrophobic chemicals like halogenated solvents, such as chlorinated ethenes and ethanes, usually do not bind strongly to the soil constituents and instead are either evaporated following emission to soil, or tend to reside in the groundwater. A specific class of chemicals that, other than those mentioned above, are intentionally emitted to the soil compartments and more or less by definition exert adverse effects, are the pesticides. Apart from their unique mode of entry (deliberately spraying) into the soil, pesticides behave like most persistent organic compounds and this class of compounds will not be dealt with exclusively in this chapter. Apart from classifying chemicals as naturally occurring, xenobiotic, essential or non-essential substances, it is also possible to classify them further on the basis of their inherent chemical properties and toxicity characteristics. POPs is a diverse category of bioaccumulative and toxic organic compounds of natural or anthropogenic origin that resist photolytic and chemical degradation,

Fate of contaminants in soil

249

C1 C1

CI C1 C1

CI~CI C1

C1 C1

C1

Aldrin

C1 ~ C1

CI

C1~

DDT

C1 ~ _ _ ~ ~ . . C1 C1N ~ N , ~

C1

CI

c1

Chlordane

7

C1

C1 C1

C1 Z,~ C1

.el

C1C I ~ ~

~CI

~C1

C1

C1

Hexachlorobenzene

Mirex

PCB

Phenanthrene

Fluoranthene

Benzo(a)pyrene

~c1 ck,~,c1

OH O

! ~ O OH

II

1

Cl C1 Dioxins

C1 Furans

HO~C~

II

r

OH

COOH

I I I

OH

y

HO~C 2,4-dichlorophenol

OH + Syringic

C1 ~-

OCH3

Humified structure

oHOH

[J~

/ / dH

H

C1

I

HO~ C ~ C ~ O

C

V

"OH

II o Humus structures

acid

Figure 3. Structures of some of the priority persistent organic pollutants, as identified by the United Nations Economic Commission for Europe but condensed according to QSARs (Quantitative Structure Activity Relationship), and a number of PAHs typically found in soils, and some humus-like structures

250

W.J.G.M. Peijnenburg

and which require specific conditions for microbial degradation. They are characterised by low water solubility and high lipid solubility, resulting in bioaccumulation in the fatty tissues of living organisms. POPs may be transported in the environment at low concentrations through the movement of fresh and marine waters. Recent evidence has shown that POPs are also transported long distances in the atmosphere, resulting in their widespread distribution around the earth even to regions where they have never been produced or used, like the Arctic. Concern about the possible long-term adverse effects on human health and the environment has initiated international negotiations on an 'International Legally Binding Instrument for Implementing International Action on Certain Persistent Organic Pollutants'. In these negotiations, twelve persistent organic pollutants have been identified as priority POPs (Table 1), requiring immediate legislative action. A selection of some typical substances is given in Figure 3. It also contains some typical substructures of humus-like soil components that make up the major part of the organic material present in soils. Most of the priority POPs have half-lives in the soil that vary from several months to several years. The half-lives given in Table 1 should be regarded as tentative because they vary with environmental conditions such as redox, temperature, and pH. Delicate interactions between chemical structure, soil properties, and mode of entry into the environment determine whether or not a specific chemical is persistent and hence of potential hazard to the soil compartment. It is the interaction with the constituents that make up the soil matrix in combination with soil-specific factors that determines whether a contaminant will be incorporated into the cell material of plants and animals, or is subjected to a number of natural attenuation processes. A schematic overview of the advective and diffusive processes governing the environmental fate of chemicals is given in Figure 4, as a more detailed example of the multitude of specific interactions typically occurring in top soil. The processes of sorption and precipitation grosso modo determine the fate of contaminants in the environment. 9.2.3 Metals and nutrients

In addition to the organic compounds mentioned above, metals and metal compounds as well as nutrients may affect the functioning of soil ecosystems. Ranges of metal concentrations typically encountered in soils, are summarised in Table 2. The table clearly shows the wide range of trace element levels that are typically found. In general the lower-end values given here represent natural background levels.

Fate of contaminants in soil

251

Table 1. Properties, the half-lives in soil, and a general description of the characteristic environmental behaviour of the priority persistent organic pollutants identified by the United Nations Economic Commission for Europe (modified after Mackay et al., 1999) POPs

Properties

Half-lives Koa

Aldrin

~ 106.5

~ 10 -1-4

~

107.9

10,000 - 30,000

Persistent

+

Chlordane

~ 106.o

~ 10 -2.9

~

108.9

10,000 - 30,000

Persistent

+

~ 106.2

~ 10 -3.2

~ 109.4

10,000 - 30,000

Persistent

+

trichloroethane

humification

(DDT)

Hexachlorobenzene

~ 105.5

Mirex

~ 106.9

Polychlorinated

~ 106.5

~ 10 .2.3

- 106.9

~

biphenyls

behaviour in soil

Ageing/

Kaw

Dichlorodiphenyl-

(h)

Environmental

Kow

~ 1 0 -1.3

~ 106.8

>30,000

Persistent

+

10 .0.5

~ 107.4

>30,000

Persistent

+

~ 108.8

>10,000

Persistent

+

2,3,7,8,-congeners

+

~

(PCBs)

Polychlorinated

dioxins

10 -2.9

~ 109.7

3,000- 10,000

especially toxic Persistent Polychlorinated

furans

~ 106.1

~ 10-3.2

~ 109.3

10,000 - 30,000

2,3,7,8,-congeners

+

especially toxic Persistent Phenanthrene

~ 104.5

~ 10 -4-9

--

2.5 - 4 , 4 0 0

Mineralization

+

Fluoranthene

- 105.2

~ 10 .5.5

--

2.5 - 4 , 4 0 0

Persistent

+

10 -6.3

--

Persistent

+

Benzo(a)pyrene

~ 106.3

Kow - o c t a n o l - w a t e r Kaw = a i r - w a t e r Koa = o c t a n o l - a i r

partition

partition partition

~

1,368- 13,000

coefficient

coefficient coefficient

Metals and nutrients are introduced into the environment from natural and anthropogenic sources and due to (lack of) specific interactions with the soil constituents. Both enhanced and diminished levels of metals and nutrients may affect soil ecosystems. It is even possible that both deficient and excess levels of metals and nutrients are present in a single soil. The complexity of balancing the levels of contaminants in the environment is well illustrated by the fact that the introduction of (artificial) fertilisers to optimise crop yield is often associated with the release of heavy metals like Cd that are present at low levels in natural phosphate ores. Interpretation of the consequences of the long-term presence of contaminants for the functioning of ecosystems and h u m a n beings must take into account the site-specific potential of natural attenuation. Differences in availability for interaction of the contaminants with plants and animals are also important. Thus, whereas nutrients in a given soil may lead to adverse

252

W.J.G.M. Peijnenburg

Fate of contaminants in soil

253

eutrophic effects, similar concentrations in soils with different physico-chemical properties may lead to deficiency. During the last fifty years in particular, excessive manure and xenobiotic compounds have entered the environment in large amounts. As far as manure is concerned the nitrogen, phosphorous and sulphur cycles have been investigated in detail as soil fertility aspects. The fate of these components in the soil will not be discussed in detail in this chapter. One aspect, the cycling of nitrogen, is presented here (Table 3). The intermediates are, in general, soluble, mobile and readily available for uptake. Excess nutrients result in eutrophication phenomena. However, where the source is eliminated within a restricted period of 5-10-25 years, eutrophication problems such as for nitrogen will fade away from the top soil relatively quickly. The phosphor problems for subsoil and groundwater are much more persistent. 9.3 Natural attenuation

The USEPA (United States Environmental Protection Agency) has defined natural attenuation (NA) as the processes of biodegradation, diffusion, dilution, sorption, volatilization and/or chemical and biochemical stabilization of contaminants, that effectively reduce toxicity, mobility or volume to levels that are protective of human health and the environment. The key elements determining the fate and hence the NA of most chemicals suspected of adverse effects are sorption and biodegradation. NA involves the characterization of the fate and transport of contaminants to evaluate nature and extent, the evaluation

Table 3. Reaction processes in the N cycles (modified after De Vries and Breeuwsma,

1987, and Wessel, 2001) Process N cycle

Reaction Equation

Reverse Process

F i x a t i o n o f N2

4 R O H + 2N2 + 3 C H 2 0

~

4 R N H 2 + 3CO2 + H 2 0

Ammonification

RNH2 + H ++ H20

~

ROH + NH

Volatilization

NH4 §

~

NH3 + H §

Nitrification

N H 4 § q- 0 2

~

NO3 + 2H § + H20

U p t a k e of N O 3

R O H + NO3- + H § + 2 C H 2 0

~

R N H 2 + 2CO2 + 2 H 2 0

Denitrification

5CH20 + 4NO3 + 4H §

~

2N2 + SCO2 + 7 H 2 0 4 N O 2 + CO2 + 3 H 2 0

Chemodenitrification

CH20 + 4NO3 + 4H §

~

A b s o r p t i o n of N H 3

ROH + NH3

~

RNH2 + H20

A b s o r p t i o n of NO2

4ROH + 4NO2 + 7CH20

--*

4 R N H 2 + 7CO2 + 5 H 2 0

O x i d a t i o n of N O 2

4 N O 2 + O2 + 2 H 2 0

--*

4NO3 + 4H §

W.J.G.M. Peijnenburg

254

of the factors that affect the l o n g - t e r m p e r f o r m a n c e of n a t u r a l a t t e n u a t i o n processes,

and

the

monitoring

effectiveness. In the N e t h e r l a n d s

of

contaminants

to

ensure

continued

this a p p r o a c h t e n d s to b e c o m e c o m m o n

practice. The c o n t a m i n a n t s that typically p o s e a threat to soil e c o s y s t e m s are either e x t r e m e l y toxic, or there is a limit to their availability for N A p r o c e s s e s a n d u p t a k e b y plants a n d animals. A schematic o v e r v i e w of the N A p r o c e s s e s that affect the persistence of soil c o n t a m i n a n t s is g i v e n in Table 4. For P O P s a n d P A H s , a n d h e a v y metals this will be d i s c u s s e d in t e r m s of sorption, leaching, biodegradation

and

plant uptake

aspects.

The c h a p t e r

of V e r s t r a e t e a n d

M e r t e n s is a p p r o p r i a t e for e u t r o p h i c a t i o n a n d n u t r i e n t cycles.

9.3.1 Sorption in NA An

overview

of the s t r u c t u r e s

of the POPs,

PAHs

and

the h u m u s - l i k e

c o m p o n e n t s (Figure 3), indicates that similar reactive sites are visible. Free e l e c t r o n - p a i r s a l o n g the o x y g e n a t o m s a n d at the e d g e s of the chlorine a t o m s provide

many

opportunities

of

co-valent

binding,

and

hence

of

tight

i n t e r a c t i o n s b e t w e e n these chemical s t r u c t u r e s a n d the c o n s t i t u e n t s of the soil matrix. S u b s e q u e n t l y , the c o n t a m i n a n t s b e c o m e tightly b o u n d to the soil m a t r i x a n d c a n n o t be easily extracted b y m e a n s of c o n v e n t i o n a l techniques, a n d are

4. Overview of the natural attenuation processes that affect the persistence of typical soil contaminants

Table

Natural Attenuation processes

Heavy metals

POPs

PAHs

Nutrients

+++ + + + ,_, ++++ +++

+++ + + _,_~ ++++ +++

++ ++ ++ + ++++ +

++

++

++++ ++

in soil

Sorption ++ Diffusion + Dilution + Evaporation Microbial transformation +1) Chemical/biochemical + stabilization Uptake plants ++ Uptake soil fauna ++ ++++ : very general and strong +++ : general and strong ++ : regularly + :seldom/hardly : not 1) : an electron acceptor (Mn, Fe, etc.)

Fate of contaminants in soil

255

not available for interactions with soil biota. The phenomenon of incorporation of initially potential toxic compounds into soils is referred to as humification, bound-residue formation, or ageing. The complexation of organic contaminants with humic material can also be called humification and is part of ageing. The binding to humus decreases the amount of material and reduces the toxicity of the parent material. Jean-Marie Bollag (1992) emphasizes that this process should be exploited. He describes for instance the incorporation of 2,4dichlorophenoxyacetic acid in syringid acid, a humic acid (Figure 3). It is an oxidative coupling process. The products formed during the cross-coupling of pollutants and humic acids are highly heterogeneous and complex and, therefore, are extremely difficult to identify. In laboratory studies some mechanisms are elucidated by using model systems: fungi producing laccase incubating with chlorinated phenols in the presence of specific phenolic humic constituents. The process of humification has been neglected or remains unmentioned in the natural attenuation process, and therefore receives some special attention here. The higher the natural organic matter content of a soil, the more feasible the phenomenon of NA by sorption to and inclusion in organic matter. The effect of ageing is schematically depicted in Figure 5 according to the idea of Fuhr (1984) who took into account that chemicals may be partly transformed and ultimately bound, leading to a bound residue. From Figure 5 it can be deduced that the (non-bound) fraction of the chemical concentration that is available for interaction with the pore water and the biota present in the soil, diminishes quickly over time. This has severe consequences for extrapolating results form short-term tests, such as those usually done in laboratory settings, to realistic field conditions. For instance, toxicity testing is usually done after relatively

100 mineralised to C O 2

(~

bound residue

0

1

2

3

5

10 time (months)

Figure 5. Schematic representation of the impact of time on the fate of chemicals in soil

(modified after Fuhr, 1984)

256

W.J.G.M. Peijnenburg

short equilibrium periods that typically vary from a few minutes to a few weeks.

Equilibrium Partitioning As explained by Jager (2003a), the leading theory on the uptake of chemicals by soil organisms is the Equilibrium Partitioning (EP) theory, formulated and broadly adopted around 1990. Basically, this approach states that organisms do not take up chemicals from soils directly, but only from the freely-dissolved phase in the pore water. Thus it may be deduced from Figure 5 that the results of toxicity tests, for instance, will typically overestimate the truly bioavailable fraction and hence adverse effects in the real environment. A chemical will tend to distribute itself between the soil, water and organism phases until it is in thermodynamic equilibrium. This implies that the chemical residues in organisms can be predicted when we know the sorption coefficient of the chemical (partitioning between solids and water) and the bioconcentration factor (partitioning between water and organism). This is schematically depicted in Figure 6. EP theory takes the effect of ageing into account in risk assessment. EP has become an integral part of soil chemical risk assessment as far as predicting toxicity (from aquatic data) as well as body residues (from total concentrations) in soil-dwelling organisms are concerned. Currently, the term EP is often used in a broader sense, relaxing the precondition of equilibrium, and denoting the fact that (time-varying) concentrations in organisms can be predicted from the

Figure 6. Schematic overview of the processes underlying the equilibrium partitioning

concept (modified after Jager, 2003a)

Fate of contaminants in soil

257

(time-varying) concentrations in pore water. Despite its popularity, the limitations of EP have also been observed. The most striking deviations are discussed below. Sequestration or ageing, as mentioned above, is the process by which chemicals tend to become less available in time for uptake by organisms and by soft chemical extraction techniques. The most likely mechanism for this behaviour is that the chemical is moving deeper into the organic matrix as contact time increases, or because of possible chemical oxidation with humic substances as Bollag suggested. Sequestration has been presented as a deviation from EP, but in fact it is a strong support. Granted, the use of equations where sorption is estimated from hydrophobicity will fail to predict the effects of sequestration, but EP (in the broad sense) appears to be quite robust as long as good estimates or measurements of pore-water concentrations are available for the specific situation being studied. Another deviation from EP that is extensively discussed is feeding. Chemicals are not only taken up by organisms from (pore) water through the skin, but also from the gut. It is a generally held view that the existence of multiple routes of entry into an organism leads to deviations from EP predictions, especially for very hydrophobic chemicals that have a low solubility in water. It was predicted that feeding becomes an important uptake route for earthworms when log Kow exceeds 5. For sediment organisms there is evidence that feeding is important for very hydrophobic chemicals, and may lead to deviations from EP of up to a factor of 5. However, there are few studies that succeed in experimentally separating both uptake routes, and often conclusions on uptake routes are drawn without confirming that equilibrium was established, and without knowing the actual pore-water concentrations either. Furthermore, it is unlikely that chemicals are transferred directly from a solid phase to an organism without the intervention of a solution phase. Biotransformation may also lead to deviations from EP, but this process is not well studied. Biotransformation usually results in more water-soluble metabolites that are easily excreted. When the exchange with the pore water is slow, even low levels of transformation may affect toxicokinetics. Despite its limitations, EP is still the reference theory for discussing the accumulation of organic chemicals in soil organisms. This implies that body residues observed in biota should first be related to pore water concentrations before alternative theories can be explored. Intermedia transport is important in systems that contain more than one phase. Chemicals will migrate from one phase to another if the phases are not in thermodynamic equilibrium (do not have the same fugacity). Octanol is often

258

W.J.G.M. Peijnenburg

considered as a surrogate for various condensed lipophilic materials present in natural phases such as non-living natural organic matter present in soils, sediments, or aerosols and certain lipid-like constituents of plants, animals, and micro-organisms. The fraction of such organic phases on a global scale is quite low. Organic phases, however, have a high affinity and a high storage capacity especially for organic contaminants. Organic phases, therefore, are the major sinks for hydrophobic contaminants. It has been observed experimentally that the ratio of concentrations in two phases is constant if the concentrations of the chemical in both phases are sufficiently low (thermodynamic equilibrium). In this case, at equilibrium conditions, the reversible distribution between phases can be described by a constant, which is known as the distribution coefficient (Kab):

Kat , = Ca

(1)

Cb For solids - (pore) water systems, the equilibrium constant is known as the partition coefficient (Kp) or distribution constant (Kd). Partition coefficients are available for many chemicals from laboratory and field measurements. As organic carbon present in water (Dissolved Organic Carbon), sediment or soil is the main sink for hydrophobic organic contaminants, the partition coefficients for these compounds are often adjusted (normalised) with respect to the organic carbon content of these compartments:

K

Cs

P

-Kocxfoc =~

Cw

(2)

Koc is the organic carbon normalised partition coefficient (L/kg), foc is fraction of organic carbon, and Cs and Cw are the chemical concentrations in the solid phase and the (pore) water phase respectively. Koc for neutral organic chemicals is often estimated from the octanol-water partition coefficient (Kow). It may be deduced from Equation 2 that partition coefficients of hydrophobic organic compounds in general are dependent upon both the chemical of interest (compound specific properties affect the value of Koc), and the properties of the medium in which it resides. Apart from the fraction of organic carbon present in the sorption phase, environmental factors also affect partitioning. These factors include temperature, particle size distribution, surface area of the sorbent, pH, ionic strength, presence of suspended material or colloidal

Fate of contaminants in soil

259

material, as well as dissolved organic matter concentration. In addition, clay minerals may also act as additional sorption phases for organic compounds. Nevertheless, organic carbon-normalised partition coefficients for a specific chemical are fairly constant among soils, provided that the additional environmental factors impacting partitioning are kept reasonably constant. Highly hydrophobic contaminants like POPs and high-membered ring PAHs tend to sorb strongly to organic phases present in the solid and the dissolved organic carbon (DOC) in the aquatic phases (pore water). Uptake of POPs and other hydrophobic contaminants therefore usually takes place directly from the solid soil phase such as ingestion of solid material and consumption of food. Cationic pollutants like ionic substances and metals show deviant behaviour. Apart from organic carbon, additional constituents of the solid soil phase may act significantly as sorbing phases. Again, environmental conditions like pH and redox state strongly affect the partitioning of these compounds and as a result Kp or Kd values vary by orders of magnitude among different soils. The effect of pH and redox conditions (Eh) is schematically illustrated in Figure 7 (see Salomons, 1995). Metal levels in the pore water will be especially high in aerobic soils at low pH (sandy soils), whereas it can easily be deduced from Figure 7 that metals will be dominantly sorbed or precipitated in anaerobic soils

Environmental processes

1.0

Changes in mobility

1.0 -Cr

m

Eh

Eh

0.6

0.6 -

n

0.2

m

Mn

~

Ni

0.2

-0.2 -

~a~lon

rise groundwaterlevel

-0.6

-0.2

-0.6

I

I

,

, pH

pH

Figure 7. Schematics of the effect of pH and Eh on metal speciation (modified after

Salomons, 1995)

260

W.J.G.M. Peijnenburg

at high pH. Hence, the latter soils in general pose no threat to organisms exposed via the pore water. The master variables, pH and Eh, affect the processes of adsorption, precipitation, complexation and oxido-reduction of the elements (heavy metals) and their solubility. Depending on soil type it can be concluded that acidification increases the solubility and therefore the mobility of the metals Cd, Pb, Zn, amongst others. Since Cr has a more extreme speciation values (Cr 3§ versus Cr 7+ in CrO4-) it may behave differently. Desiccation, leading to more oxidizing conditions, may enhance mobility. Changing the groundwater level results in other environmental conditions and therefore other mobilities. Speciation of heavy metals in the aqueous phase, including sorption to DOC, strongly affects the availability of the metals for uptake and subsequently exerts adverse effects. Speciation is usually associated with the aqueous phase and involves the interaction of a contaminant with the various organic and inorganic constituents of the pore water. In the case of metals, these interactions not only include the formation of short-living complexes with a number of anions, like chloride, sulphate or carbonate, and complexation or binding to DOC, but also binding to the outer membranes of organisms and plants. An example of the various interactions in pore water is given in Figure 8. In the same way that soil organic matter is important for sorption of hydrophobic organics, the sorption to DOC is an important attenuation process for these

Figure 8. Schematic overview of the interactions

of a metal ion in solution

Fate of contaminants in soil

261

chemicals. They effectively reduce the amount of chemical present in the pore water that is available for sorption to living biota.

9.3.2. Leaching Leaching is of relevance for all chemical substances that may dissolve in pore water or groundwater. Leaching is the advective movement of a pollutant towards the groundwater. Leaching is directly related to the total concentration of the contaminant in the pore water, including the fraction sorbed to the DOC. The hydrophobicity of an organic contaminant in combination with the amounts of organic carbon in the solid phase and the pore water determine leachability in a given soil. Leaching of metals is also directly proportional to the total metal levels in the pore water and is strongly affected by the soil and pore water properties that determine metal partitioning in the soil. Especially pH is an important parameter in this respect. As such, the leachability of a chemical is directly related to a delicate balance that results from the sorption and speciation processes discussed above. In general, higher concentrations of particulate organic matter and higher pH-values decrease the extent of leaching of any contaminant, except Cr and As. The Henry constant (H) of a chemical determines the volatilization. This is the air-water partition coefficient (Kaw). It can be expressed as a dimensionless ratio of the concentration of the chemical in the air and in the (pore) water. For any organic compound, it is the magnitude of the Henry's law constant in relationship to the value of the Kocthat determines the extent of volatilization of the chemical. A multi-media fate model is an appropriate means for describing the extent of volatilization of a given contaminant in a given soil.

9.3.3 Volatilization in NA Volatilization is relevant only for some POPs and PAHs. Volatilization may contribute directly to the attenuation of contaminant levels in soil. Many substances reach the soil as a result of wet and dry deposition from the air. Whether a volatile chemical will partition from the soil into the air depends on its interaction with various soil constituents. Thus, volatile hydrophobic chemicals like a number of PCBs that sorb strongly, for instance, to organic carbon will only volatilise in limited amounts.

9.3.4 Biodegradation and metabolism in natural attenuation Biodegradation is mainly important in the case of organic contaminants. Biodegradation, or the transformation of chemical substances through the action of living organisms, is one of the major processes that determine the fate

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of organic chemicals in terrestrial environments. If the biodegradation of a chemical is slow, then the chemical may bioaccumulate and cause primary and secondary poisoning in the food web, or it may reduce the quality of drinking water and affect the various functions of surface waters. Biodegradation is, in general, equivalent to conversion into simple and relatively harmless molecules and ions, such as carbon dioxide, methane, water and other inorganic compounds. This phenomenon has been called ultimate biodegradation or mineralization and may be regarded as a true sink in aerobic soil compartments. The environmental conditions of the soil are the key factors that determine whether transformation and mineralization occur. Even in arable soil systems, soil aggregates host a variety of different redox-conditions and, therefore, it is plausible that the potential for degradation is there for most, if not all, organics. The degradation itself will be slow as Table 5 shows. As can be seen in Table 5 natural degradation rates of plant material are slow (k is the relative fraction degraded per time unit). Those rates of natural organics are presented in order to be aware of the extreme slowness of normal natural biodegradation and also as a warning and contrast for degradation rates measured under laboratory conditions. Degradation rates of mono aromatics such as benzene under field conditions are of similar value (Hutchins, 1991; Beller et al., 1992). Natural organic matter consists of carbon, hydrogen, nitrogen, oxygen, phosphor and sulphur, the same basic material as microbes. PAHs consist of carbon, hydrogen and oxygen and POPs of carbon, hydrogen, oxygen and chloride. So, for the degradation of PAHs and POPs the microflora needs at least extra nitrogen, phosphor and sulphur, and it has to adapt its degradation system. Degradation rates of hydrophobic PAHs and POPs are extremely slow, as the response of the microbial community towards natural as well as Table 5. Some typical examples of degradation rate constants (k) of organic matter of

defined vegetations (modified after Werff, 1992) Type of predominant vegetation

Phragmites australis Phragmites karka Typha domingensis Typha glauca Typha Latifolia Typha angustata Typha elephantina Scirpus,fluviatilis

k

0,0035 0,0045 0,0078 0,0014 0,0043 0,006 0,0038 0,0018

Type of predominant vegetation

Scirpusamericanus Scirpusmucronatus Juncussquarrosus Juncusroemerianus Paspalumrepens Carex rostrata Carexriparia Zizaniaaquatica

k

0,0021-0,0025 0,0044 0,0013 0,0016-0,0017 0,00717 0,0046 0,0029 0,077

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xenobiotic organic compounds does not depend on the total concentration of the contaminants but mainly on the water soluble concentration. This soluble fraction is relatively small for mineral oil, PAH, PCB and many chlorinated organic compounds such as hexachlorocyclohexane (HCH) and HCB, but not for tetrachloroethene (PCE) and trichloroethene (TCE). The sorbed or bounded fractions of xenobiotics are not available for microbial interference as microbial activity takes place in the soluble phase. The soluble concentration may be too low to support the measurable growth of the microbial population. However, even without growth micro-organisms may completely degrade pollutants. A brief review of literature data on persistent soil contaminants as PAHs and some POPs shows that PAHs are mainly biodegraded under aerobic conditions, but the smaller hydrocarbons also with NOB or Fe203 as electron acceptor (Cerniglia, 1992; Warith et al., 1992). HCH can be degraded under many different redox conditions (Bachman et al., 1988). PCBs are biodegraded under sequential anaerobic and aerobic conditions (Quensen et al., 1988, Abramowicz, 1990). Not much has been reported on complete degradation of POPs as drins and HCH, although one may question whether this has been investigated at all. Scientific research in the field of microbiology carried out during the last two decades has lead to great discoveries. Twenty years ago the anaerobic microbial degradation of halogenated compounds was generally considered as "not occurring". In the 1980s, reductive dechlorination (Bouwer and McCarty, 1983) was discovered as a new phenomenon. Then around 1992, partly as a result of the PhD theses of Oldenhuis, Van der Meer and Holliger, the facts and principles of the microbial degradation of chlorinated compounds became more clear. The anaerobic microbial degradation of chlorinated xenobiotics also became general knowledge, and applicable to bio-remediation. It became clear that bacteria could use PCE and other chlorinated compounds as electron acceptors. The principle of "halorespiration" was added to the phenomenon of recycling of elements. Suitable electron-donors and suitable environmental conditions became objects to study to apply dechlorination in insitu bioremediation processes. It took 12 years before the principle of biodegradation of HCH was partly unravelled. In addition, the role of the various prevailing electron acceptors, such as 02, NO3-, Fe3§ 8042- and CO2 and various electron donors became clear. For example, the fact that in the presence of nitrate complete dehalogenation is impossible. Around 1995, the partial biodegradation of dioxins (Adriaens et al., 1995; Halden and Dwyer, 1997) was described and in 2003, the complete reductive dehalogenation of dioxins by the bacteria Dehalococcoides was reported (Bunge et al., 2003). Many parameters affect the biodegradation of chemicals in the environment

W.]. G.M. Peijnenburg

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as well as in biodegradation test systems used to simulate the environment. The bioavailability of the chemical to the enzyme has been identified as a major factor in determining biodegradation in nature, and this aspect will be dealt with in more detail in Section 9.5. The variables that potentially affect rates of biodegradation can be ranked into three categories: the soil related, the microflora related and the contaminant related factor, and are depicted in Figure 9. For detailed information on this triangle the chapter by Verstraete and Mertens is very appropriate. Figure 9 is presented here as an important reminder. Since the impact in soil of most of the variables given in Figure 9 is poorly understood, it is currently impossible to accurately predict how rates of biodegradation might vary from location to location for a given set of environmental conditions. Moreover, the use of second-order kinetics to evaluate rates of biodegradation has been validated only for a limited range of organic compounds. Transformation of chemicals in the environment can also occur by abiotic processes. The most important abiotic transformation processes can be divided into four separate categories, such as hydrolysis, alteration of the chemical structure by direct reaction with water; oxidation, a transformation process in which electrons are transferred from the chemical to a species accepting the electrons; reduction, the reverse of oxidation, electron transfer takes place from a reductant to the chemical to be reduced; and photochemical degradation, transformation due to interaction with sunlight. Usually, this last process is of limited relevance for the soil compartment. Transformation and mineralization processes can alter physico-chemical and toxicological properties and reduce exposure concentrations of environmental

Substrate related

Figure 9. Variables known to affect biodegradation rate constants (k) of organic

contaminants

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chemicals. Where biotransformation is performed by higher organisms, the formation of polar transformation products (metabolites) can also provide an important method of detoxification. The rate of degradation of a specific chemical depends on its availability for reaction, its intrinsic reactivity, the availability of the reactant and the reactivity of the reactant. Biodegradation and metabolism of heavy metals by microbes does not occur in a strict sense in natural attenuation since these processes imply elimination. The microbial transformation of metallic minerals, however, has been known for hundreds of years. Bacteria are capable of changing the valences of metals and metalloids. The valences of these metals as well as their major microbial transformation reactions are given in Table 6. Knowledge obtained over the last 25 years may still follow this relatively old table. Fe and Mn do play a role as electron acceptor in the degradation of aromatics (Sumners and Silver, 1978), thus inducing natural attenuation of these chemicals.

9.3.5 Uptake as part of NA Uptake by plants and animals is, in general, the consequence of a large number of competing processes, both in the aqueous and in the solid phase, as well as at the interface between the biota and the pore water. In the definition of NA by USEPA (1993), the interaction with the living other part of the soil, such as roots and soil fauna, has not been mentioned. Those living parts may nevertheless play a role in the elimination of contaminants, both by means of degradation and by accumulation. Hyperaccumulating plants may be especially important in this respect as they are capable of trapping large amounts of soil contaminants. Vegetation plays a major role for nutrients and nutrient cycling in the soil, since most plants have optimized their modes of uptake and elimination of nutrients in order to cope with conditions of both deficiency and excess nutrients. Vegetation may also play a role for the attenuation of heavy metals, albeit that this is usually on a smaller scale than nitrogen, for example. Nutrients and

6. Micro-organisms mediated heavy metal oxidation-reduction reactions (modified after Summers and Silver, 1978)

Table

Transformation

Metal

Reduction

As s*, Cr 6§ Fe 3§ H g § Hg 2§ M n 4§ Se 4§ Te 4+

Oxidation

As 3§ Cr 3+, Fe o, Fe 2§ M n 2§ Sb 3+

Methylation

As 5§ Cd 2+, H g 2+, Se 4§ Sn 2§ Te 4§

Demethylation

H~ 2+

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W.J.G.M. Peijnenburg

heavy metals, such as cations or anions, are taken up by all plants via the pore water. Roots produce or excrete organics as root-exudates. These compounds may strongly affect the availability of contaminants. Figure 10 shows the role of roots. Uptake by roots of organics such as POPs and PAHs is, in general, unnatural. Exceptions occur as the result of damaged roots and organics that behave as cations or anions. Although there are exceptions, it may be stated that uptake of POPs and PAHs via plant roots and their subsequent transport to the above ground parts of the plants is of limited importance for leafy crops. For these contaminants and these plants it is atmospheric deposition on the leaves and subsequent transport through the cuticle that determines accumulation. The uptake of vegetation may be stimulated as well as hindered, or protected, by its mycorrhiza. The role of root exudates is manifold, such as increased availability of contaminants and a higher microbial activity, called the rhizosphere effect. The chapter by Ernst on vegetation and organic matter provides more information.

9.4 Bioavailability and bioaccessibility Due to the tendency that available fractions of contaminants in soil are considered to be far more relevant than the total fraction, extensive attention will be paid to the determination of their bioavailable and -accessible fractions. Persistence affects the potential for NA and environmental exposure because persistent chemicals exhibit higher concentrations per unit emission. A key role in the processes governing the persistence and toxicity of contaminants is

Figure 10. Schematic overview of the interactions at the root-soil-pore water interface determining uptake of nutrients and metals in particular

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played by the phenomena of bioavailability and bioaccessibility. Bioavailability may be defined as the fraction of the total amount of a chemical present in soil that, within a given time span, is either available or can be made available for uptake by (micro)organisms or plants, from either the organism's direct surrounding or the plant, or by ingestion of food. This definition of bioavailability implicitly suggests that the 'bioaccessible fraction' is defined as the fraction of the total amount of a chemical present in ingested soil particles that, at maximum, can be released during digestion. Different factors affect bioaccessibility. For example, the bioaccessibility of metals and ionizable contaminants is expected to be highly dependent on the pH values in the different compartments of the gastrointestinal tract. As a consequence, the bioaccessible fraction of ionizable contaminants and metals present in ingested soil is, in general, larger for mammals than for soil-dwelling invertebrates like, for instance, earthworms that regulate gut pH at around neutral. Uptake of chemicals involves the passage of compounds across a biological membrane, mediated by a carrier or a single solute. Compounds may enter tissues through passive diffusion, facilitated diffusion and by active transport mechanisms. Passive diffusion is the major uptake process for many organic chemicals as well as some metals and organometals. The driving force for uptake is a fugacity difference between water and the organism, as explained on the basis of EP theory. The fate and toxic potential of metals is, however, governed by more refined interactions at the micro level. Ultimately it is the free metal ion that is supposed to be capable of traversing biological membranes. As a consequence, metal availability and toxicity are functions of water chemistry, since speciation determines the free metal ion activity. It is the free metal ion concentration that often provides a better indication of availability and toxicity (Sunda and Guillard, 1976; DiToro et al., 2001) than total or dissolved concentrations. Studies with aquatic organisms (Campbell, 1995; Hare and Tessier, 1996), invertebrates (Kiewiet and Ma, 1991) and plants in nutrient solutions (Lexmond and Van der Vorm, 1981) revealed that metal uptake is also influenced by protons (H § competition and divalent macro-ions such as Ca 2§ as illustrated earlier in Figure 8. Morel (1983) and Pagenkopf (1983) introduced models in which the interactions of chemical species with organisms are used to predict trace metal uptake and toxicity. Most commonly this approach is called the free ion activity model (FIAM). The FIAM is gaining popularity in studies of soilplant relationships (Parker et al., 1995), even though some exceptions are known to exist (Campbell, 1995). Recent research (Gorsuch, 2002), has advanced

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the current level of understanding of metal availability to aquatic life via aqueous exposures, and the understanding of how to relate exposure levels to effects. The relationships between chemical forms of metals and the bioavailability of the metals are not sufficiently understood for terrestrial plants and soil-dwelling organisms, to permit the prediction of metal toxicity on the basis of the chemistry of the pore water. The examples given conceal a typical, fundamental difference between exposure in aqueous systems and in soils. Terrestrial plants may well be exposed in aqueous nutrient solution. Various data have been reported which either make clear the validity of FIAM, or from which additional chemical species that are actually taken up may be deduced. These species include complexes of metals, chelating agents, natural organic ligands, such as humic acids, and inorganic ligands such as chloride. There are very few studies with soil-grown plants, and often it is difficult to separate the actual ionic species that contribute to enhanced uptake. In addition, the soil itself exerts a diffusion limitation to metal transport because of the strong nature of metal binding to the solid surface and the tortuous nature of soil pores. In soils, the relationship between free metal activity and metal uptake is, therefore, not as close a relationship as had previously been assumed. It can be concluded that membrane or dermal uptake of contaminants can be described by the EP theory (Shea, 1988). However, regulated uptake of nutrients and essential elements is different and also the uptake of contaminants via the digestive tract of soil biota representatives. Also organisms responding to other environmental conditions may have different effects on uptake. This is shown for earthworms in Jager's (2003b) thesis. For example, plants will physically access more metal in the soil when root growth rates are high, so that temperature and nutrition may affect plant metal uptake directly through their effects on plant growth (McLaughlin, 2001).

9.5 Monitoring techniques for bioavailability and bioaccessibility Monitoring implies repeated measurements in time at fixed places. 9.5.1 General

The equilibrium-partitioning concept provides a proper means of operationalizing the definitions of bioavailability and bioaccessibility. The liquid phase, the particulate phase and the biotic phase play leading roles here. Equilibration processes are assumed to take place between all phases. The 'biotic phase' consists of a variety of species, each with a characteristic set of exposure routes and specific characteristics. They include relatively long

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equilibration periods as compared to equilibration periods for most physicochemical processes, concentration-dependent occurrence of toxic responses, and the role of some equilibria in determining toxicant uptake. The challenge of monitoring bioavailable and bioaccessible fractions thus boils down to linking chemistry in the solid phase and in solution, to uptake by biotic species. It determines whether or not toxic effects will actually occur. This affects the choice of extractant used to mimic bioavailability. Since there is not a single bioavailable or bioaccessible fraction, the terms potentially and actually available fractions have been introduced to enable the inclusion of the aspect of time. Any monitoring strategy to be developed should aim at providing estimates of the fractions that are potentially and actually available or accessible. 9.5.2 A s s e s s m e n t methods

Methods developed for measuring (mimicking) bioavailable and bioaccessible fractions can be grouped into three categories. Methods for assessing contaminants in (pore) waters - in the case of metals this includes an assessment of speciation and the activity of the free metal ion; single (solvent based) and sequential extraction methods, and methods with rigorous digestion procedures to determine total concentrations, including selective oxidation of organic matter and thermal desorption. The methods for assessing contaminants in (pore) water may provide an estimate of the fraction actually available and accessible, whereas extractions and digestions may provide estimates of potentially available and accessible fractions. Nutrients The search for chemical methods to determine the concentration of individual plant-available nutrients in agricultural soils began several decades ago. As a result of research focussing on estimating the quantities of nutrients that should be added to soil to achieve maximum crop yield, chemical methods were derived that provide reasonably predictive data of the bioavailability of inorganic ions necessary for plan development. Soil water can be analysed directly to determine the fraction of phosphate actually available. Certain extractants can then be employed to determine the phosphorus fraction potentially available. It is possible to extract specific mineral fractions of phosphate, and total phosphorus concentrations are determined after complete digestion with strong acids.

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W.J.G.M. Peijnenburg

POPs Various extraction methods have been developed to estimate the (bio)available fraction of contaminants. These include Tenax-beads to extract the contaminant pool resulting in a multi-phase desorption profile of which the first, fast kinetic phase is considered the bioavailable fraction. Solvent extractions resulting in less vigorous extraction than total extraction with organic solvents and solventwater mixtures have been applied and related to uptake of PAHs and atrazine in earthworms. Results of these studies show an ageing effect for both solvent extraction and the uptake of these chemicals by the organisms. Besides solvent extractions, extractions with passive samplers such as C18-disks and semipermeable membrane devices (SPMDs) have also been used. Solid phase micro extraction (SPME) is another passive sampler that has been used in water, sediment and soil samples for some time now and SPME-derived pore water concentrations have been correlated to body burdens of contaminants in biota (Van der Wal, 2003). Metals The analytical determination and chemical equilibrium models exist for determining metal speciation in solution. Models are based upon experimentally derived parameters, and models cannot be any better than the assumptions and data on which they are based. Computational procedures, based on thermodynamic principles, allow the equilibrium speciation of a system to be calculated once total component concentrations are known. Various techniques can be used to measure free metal activity in solution. The most direct method for determining free metal species is possibly through ionselective electrodes (ISEs). Voltammetric methods such as differential pulse polarography (DPP) and differential pulse anodic stripping voltammetry (DPASV) are sensitive electrochemical methods and can be used to determine the concentrations of 'labile' metal. Exchange resins are also used to measure the free metal activity or the relative lability of metals (Beveridge et al., 1989). Donnan dialysis membranes also can be used to determine the speciation of solutions. Minnich and McBride (1987) have for instance used this technique to determine the free Cu2+-activity in Cu-salts and sewage sludge-amended soils. Recently, Temminghoff et al. (2000) developed the Wageningen-Donnan Membrane Technique for directly measuring metal activities. A separation technique via Donnan equilibrium across a negatively charged ion-exchange membrane was adapted for this purpose. Diffusive gradients in thin films (DGT) have been used to assess bioavailable forms of metals. The technique utilises a gel layer to remove ions from solution. This establishes a gradient. A

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resin at the base of the gel arrests the diffusion of ions. The DGT technique is s u p p o s e d to allow free diffusion of labile species of metals (Zhang et al., 1998; Davison et al., 2000). Finally, information on metal speciation m a y be gained by separation and/or exchange methods. For instance, Gregson and Alloway (1984) used gel p e r m e a t i o n c h r o m a t o g r a p h y to separate species of Pb based on molecular weight. Extractions represent an intermediate position b e t w e e n total digestions and (pore) w a t e r samples. Research concerning metal b e h a v i o u r in soils has focused m a i n l y on the use of various chemical extractions to describe forms of metal present. Generally speaking, six types of extractions m a y be distinguished (Table 7). In addition to single extractants, sequential extraction is often p e r f o r m e d to differentiate the metal fractions present in the soil. A w i d e l y u s e d version of a sequential extraction p r o c e d u r e was first formulated by Tessier et al. (1979). An example of a sequential extraction scheme is s h o w n in Table 8. The m o s t c o m m o n approaches to total soil metal estimates aim at

Table 7. Overview of commonly used extraction techniques for extracting heavy metals

from soil Extraction type

Examples of reagents used

(Weak) salt extractions

CaCl2, Ca(NO3)2, NH4Ac, NaNO3, Mg-salts, BaC12,in concentrations from as low as 0.001 M and up to 1 M salt solutions sodium-ascorbate, hydroxylamine-HC1, sodium dithionite acetic acid, citric acid DTPA-TEA, EDTA, NTA HNO3, HC1, 'double acid' (HCI + H2SO4) Ammonium oxalate-oxalic acid, Mehlich III (dilute acid, salt and EDTA)

Reductive extractants Weak acid extractions Strong complexation methods Dilute strong acids Combined extractants

Table

8. A general Tessier scheme of sequential extraction (Tessier et al., 1979)

Extractant

Fraction obtained

1 M MgCI2or I M NaOAc 1 M NaOAc+HOAc Reducing extractants like Na2S204 (hydrosulfite) + Na oxalate, or Na-hydroxylamine.HC1 Oxidizing agents HNO3+ H202, later NH4OAc to prevent readsorption Digestion HF+HC104

exchangeable fraction bound to carbonates bound to Fe-Mn oxides, or easily reducible bound to organic matter residual fraction

272

W.J.G.M. Peijnenburg

determining total recoverable metal. Many total metal digestions involve hydrofluoric acid additions to the digest to break down Si crystal lattices and extract metal bound or trapped in these crystal structures. Total recoverable metal digestions are typically designed to stop short of destroying crystal structures. Soil digestions also usually employ elevated temperatures. In recent years microwave digestion methods have gained popularity because of increased safety and shorter digestion times. Three USEPA microwave digestion procedures exist and include HNO3, HNOB-HC1 and HNOB-HC1-HF respectively (USEPA, 1997). USEPA states that the first two procedures are designed to extract metals that can, potentially, become environmentally available.

9.5.3 Linking chemistry to biology Element uptake is controlled by chemical availability in solution as well as the capacity of the soil or labile fractions present in solution to supply that element. Most plants and biota will accumulate many times the amount of metal available in solution at any given moment. In effect the soil solution is emptied and replenished. Uptake of contaminants is, therefore, not only dependent on the availability of the chemical in solution (intensity factor) and the uptake mechanisms, but also on the capacity of the soil solid phases to supply the fluxes of that particular element (capacity factor). Understanding bioavailability for species exposed via the pore water requires the consideration of both aspects: the intensity of exposure through the EP-concept or the FIAM, and the capacity of the soil to maintain this level of free contaminant in solution. The consequence for daily practice for risk assessment, the approach can on a regular basis only be applied for the aqueous compartment of ecosystems. At present, several programs directed at incorporating further refinements into the BLMs for copper, zinc and silver are in progress in order to extend the application of these BLMs to other types of organisms, thus making the BLMs more suitable for use in the development of water quality criteria. Application of the BLM framework to soil biota is in the early stage of development studies. An elegant recent approach for measuring effective soil solution concentrations and metal supplied from the solid phase, is the technique of diffusive gradients in thin films (DGT). This technique of surrogate chemical measurement, first reported by Davison and Zhang (1994), has promise as a quantitative measure of the effective bioavailable metal for plants, as measured DGT fluxes of Cu relate well to Cu uptake by Lepidium heterophyllum (Zhang et al., 2001). A second general approach towards relating available and bioavailable fractions to uptake is by means of multifactorial analyses. In this approach,

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accumulation characteristics like uptake rate constants, steady-state concentrations, bioconcentration factors, and biota to soil accumulation factors are statistically associated with water or soil characteristics such as the total concentration, rapidly or slowly desorbing contaminant fractions, pH, organic matter content, dissolved organic carbon content, and the cation exchange capacity (CEC). Observations are typically analysed by multiple regression analyses, to weigh the impact of these characteristics against each other. It should be noted that this general approach does not necessarily provide any insight into mechanistic influences of any of the variables mentioned on metal uptake by plants or organisms. Nevertheless, such approaches allow eliminating uptake routes that do not contribute significantly. Thirdly, instead of total soil concentrations, operationally defined extraction techniques have been used with the aim of assessing the pool of solid-phase contaminant that buffers the solution concentration, thus mimicking the concentrations of labile contaminants that are available for uptake by biota or plants, and accounting for the supply rate term capacity factor. Various successful examples have been published. Often there are large discrepancies among extraction techniques with regard to simulating bioavailable and bioaccessible fractions. Whereas one specific extraction technique may provide a good indication of the bioavailable or bioaccessible fraction for a specific species or plant, there may not be any correlation at all for other species or plants using the same extraction technique. In addition, when applied to soils with a broad spectrum of contaminant concentrations or soil characteristics, they often fail to adequately predict metal availability or accessibility. Therefore it is clear there is no single expression for bioavailable or bioaccessible fractions. 9.6 Towards a monitoring strategy

The findings summarised above illustrate that the consideration of an EP approach as the basis for monitoring is an overly simple approximation of reality. A large number of simulation approaches for bioavailability and bioaccessibility are currently available and validation of the chemical techniques by biological uptake and toxicity experiments yield variable results. Whereas the paradigm of the free ion in solution in the case of metals or the non-complexed and truly dissolved POPs and nutrients as being the available species, was confirmed for some organisms or plants, other organisms or plants do not adhere at all to this paradigm. Instead, experiments with the latter biota clearly warrant the use of multivariate approaches and the plethora of extraction techniques developed so far. Although there is a large variance in species-specific, compound-specific and compartment-specific bioavailable and

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W.J. G.M. Peijnenburg

bioaccessible fractions, determining the most available or accessible pools (while not at all perfect) will provide a better prediction of risk than total concentrations or contaminants extracted by strong extractants. However, one should be aware that especially indices of actual availability or accessibility can change over time. This may be due to long-term changes in pH, redox conditions or organic matter content. Indices of potential availability or accessibility will respond less, and hence any monitoring strategy should take potentially available fractions into account. Total concentrations or compounds extracted by strong extractants provide a worst-case estimate of possible long-term changes and should therefore be taken as a starting point for monitoring, irrespective of the goal of monitoring. Before designing a monitoring strategy, it is of utmost importance to have clarity about the purpose of the monitoring programme. Issues to be considered include the composition of the ecosystem of interest in terms of environmental properties (like OM, DOC, pH, redox condition, etc.), the end point of determination (ecosystems versus specific populations, species or even individual organisms or plants), the relevant uptake and exposure routes as well as the chemicals of interest. In common practise such information will often be lacking, and instead of focussing on actually available or accessible fractions, the focus will therefore be the potentially available or accessible pools. A general monitoring strategy is depicted in Figure 11, and further elaborated in Table 9. Depending on the pre-set end point of determination and the critical total levels related to this end point, the first step of the monitoring strategy provides for a worst case approach. This can be used to deduce whether a worst-case estimate of the potentially available and accessible pool of contaminants is likely to be indicative of adverse effects. The limit indicated in Figure 11 against which the measured total content of contaminant of the potentially available contaminant concentration is to be judged, can be many. Examples include any environmental quality objective that has been set for the chemical of interest, any adverse effect level considered to be relevant for the specific site of contamination, and any effect level that is related to the objective of protecting individual species (like for instance endangered soil organisms or plants), populations, distinct parts of ecosystems, or human beings. In the case of the effect level against which contaminant concentration data are to be compared, it is preferable to use chronic effect data, rather than acute effect data. If the pre-set end point of determination is exceeded, then it must be determined in the next step whether the composition of the matrix of interestin terms of properties affecting potential availability or accessibility - will

Fate

of contaminants

in

275

soil

Table 9. Monitoring table Components .........................................................

~otal content, including bound iSoil E !Extractable contaminant content iresidues ~roper...................................................... i........................................................................................................................................................................................................................... iMethods iEnd point of ~ies of iExamples of suited iEnd point of available !consideration ~nterest methods !consideration ~.. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

~ .............................................................................

~. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

, ...................................................................................

~. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

POPs and PAHs

iHarsh ~ n y quality ..C)M -Tenax-extraction ~ n y adverse i Wextraction, ..~bjective or adverse j-Less vigorous solvent effect level like ~ffect level based on extractions based on ~rolonged ~otal concentrations, i-Extractions with C18- extractable isoxhletjIn addition, total idisks and SPMDs iconcentrations ~xtraction natural background i-SPME !i i levels are to be taken i ii ~nto account i ii . ~henever relevant i ......................................................... i ..................................................... i ............................................................................ i................................... ~................................................................................... i............................................................. Heavy metals Digestion, by ~ n y quality pH, OM, -Weak salt extractions .--~ny adverse means of !objective or adverse klay -Reductive extractants effect level istrong HNO3, ieffect level based on ~ontent, -Weak acid extractions based on JHNOB-HC1, ~otal concentrations. ~h i-Strong complexation extractable HNO3-HC1- iIn addition, total methods concentrations, 'i i IHF hatural background i i-Dilute strong acids ~sing extraci levels are to be takenl -Combined extractants tants of similar i ~nto account i iextraction i i istrength Nutrients Harsh ~ n y quality ~H, ~sually not relevant -.-...".extraction or iobjective or adverse ~OC9 i idigestion effect level based on Content iotal concentrations, k)f pore Total natural back- .~ater, i ~round levels are to ~h, OM , taken into ! be i iaccount. !

.........................................................

Eh

" ......................................................

~ ............................................................................

i .......................................................................................................................

~. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

= redox potential

OM = organic matter

sufficiently reduce the potential adverse impact of the total pool. If it is judged that the potential adverse impact is not sufficiently reduced by the composition of the soil matrix, then a large battery of experimental techniques for assessing the various contaminant pools may be applied. Depending on the end point under consideration and the compound-, matrix-

W.J.G.M. Peijnenburg

276 < risk limit [ Total content

Stop

1 risklimitl

1

< risk limit[

Physico-chemical matrix properties

Stop

11 nsk'imitI Extraction matrix

Analysis pore water; Modelling speciation

Species specific matrix

Figure 11. General outline of a monitoring strategy supporting risk assessment

and species-specific factors identified above, the proper set of measurement techniques is to be selected in order to make sure that site-specific impacts are adequately taken into account. 9.6 C o n c l u d i n g remarks

Assessment of the potential and actual adverse effects of contaminants on soil ecosystems requires that the characteristics of both the soil environment and the contaminants are taken into account. A simplified strategy that takes these elements into account is to combine expressions of total amounts of pollutants present with expressions of bioavailable fractions. For the three classes of contaminants considered combinations of commonly available methods of analysis are recommended in the integrative assessment of soil quality and soil functioning. For POPs (including PAHs) the soxhlet extraction to determine potential risks, combined with tenax extraction to assess actual adverse effects. For heavy metals it is recommended to combine HNOg-extraction with a weak salt extraction (0.01 M CaCh). The pH of the soil, OM-content, clay content, and natural background levels need to be determined in order to make sure that expected adverse effects are not merely related to the composition of the soil, or to the varying ecological preferences of the species present in the undisturbed soil ecosystem). In the agricultural context many methods have been developed to assess the nutrient status. HNO3-extraction will be sufficient to assess the potential impact of either increased or decreased levels of nutrients. Suitable reference soils should be included in the monitoring programme.

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Monitoring results of potential and actually available fractions that are m u t u a l l y supportive, imply that a decision on the next steps in terms of more detailed risk assessment can immediately be taken. More detailed research, including expert judgement, is required where the expressions of potentially and actually available and accessible contaminants give mixed results.

9.7 Implementation in soil management The final question to be a d d r e s s e d is w h a t are the consequences for practice. As far as the nature and b e h a v i o u r of contaminants are concerned it boils d o w n to the ultimate fate of the soil. Similar processes are also involved in areas called "City Green" (small green areas within the city) and "Ecological Main Structure" (large green locations outside cities). For practical purposes the factors that control contaminant behaviour have to be monitored. The fate of nutrients, heavy metals and persistent organic contaminants in soil have been described and explained in relation to recycling possibilities and soil characteristics such as organic matter, mineral fraction and environmental conditions such as p H and redox-conditions. These controlling factors can be quantified on a routine base, but they can also be modelled. Examples here include a model for extractability, bio availability for e a r t h w o r m s or bio-availability for moles, provided the uptake procedure of respectively e a r t h w o r m s and moles is included. In the foregoing it is shown that the knowledge, data and key characteristics on c o m p o u n d behaviour are available for application in soil m a n a g e m e n t , provided the specific equilibrium partitioning pitfalls have been considered. The implementation itself is the responsibility of the one asking the questions: W h a t is it you want? A n d of course there should be validation and verification of the model outputs, as unexpected differences can be better analysed in the field than from behind a computer screen.

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