assessment of the current exposure of the british natural

Assessment of the current exposure of the British

natural environment to silver (CB0464)

Silver in sewage sludge, soil and river bed-sediments

Report for Defra 6th

June 2014

(amended 5th August 2014)

Monika D. Jürgens, Andrew C. Johnson, Alan J. Lawlor, Sarah A. Thacker,

Richard J. Williams, M. Gloria Pereira, Catherine L. Barnett, Nick A.

Beresford

Centre for Ecology and Hydrology

HSAC1419_Final report of CEH Nanosilver river study` last saved 03.09.2014 13:27:00

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Executive Summary

This study found 3.56 mg/kg dw (median) silver (Ag) in sludge cake from 9 separate sewage

treatment plants (2.77 mg/kg dw as the 10%ile and 14.3 mg/kg dw as the 90%ile). Thus, perhaps

rather surprisingly, despite its status as a precious metal, the UK is disposing of approximately 4

tonnes of silver (Ag) down the drain each year. There were no dramatic differences according to

type of plant or location in the country. However, one plant had almost ten-fold more Ag in its

sludge than the others and this may be linked to the amount and nature of its industrial waste

received. Following typical agricultural recommendations on the amount of sludge that may be

applied as fertiliser to agricultural land and using our median sludge Ag content ploughed into 20

cm of topsoil would result in 11.3 µg/Kg concentration of anthropogenic Ag in soil. This addition

would represent approximately 4% of background levels. There are still only few studies on the

ecotoxicity of ionic and nanosilver in soil. The lowest effect concentration (EC50) where 50% of the

test animal died in soil studies in the European NanoFATE project was reported as 4,800 µg/Kg for

nanosilver and 6,200 µg/kg for ionic silver. Within the NanoFATE project a predicted no effect

concentration of 1,240 µg/Kg for ionic Ag was proposed. Thus, according to the calculations in this

study, a single sludge application to soil would introduce anthropogenic Ag at a level 100 times

below a possible PNEC for ionic Ag.

Using freeze coring to examine river bed-sediments in the River Thames revealed that in rural

locations, with modest human input, typical Ag values were 0.02-0.04 mg/kg dw. However, at sites

with a high human/industrial input, values of 1-5 mg/kg dw were found. This compares to

background silver measurements in topsoil recorded in the European geochemical baseline

database of a median of 0.3 mg/Kg. Thus, with the exception of the site immediately downstream

of Oxford’s Littlemore sewage works, all sediments measured were in the same range as the

background levels recorded for soils. Some sediment layers from that site however had even more

Ag than most sewage sludges! This indicates that exposure of river benthic organisms to Ag at

some sites may be greater than other wildlife in the UK. In our previous study carried out for Defra

colloidal Ag was around 10% of total Ag discharge in sewage effluent and in a study carried out on

German STPs this was around 5%. Thus, nanoparticulate Ag would appear to be only a small part of

total Ag discharge to the aquatic environment so far. Using Cs-137 to date the sediments it was

found that they were, for the most part mixed, making it difficult to state whether Ag entering river

bed-sediments is going down or up.


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This study appears to raise somewhat our concerns about Ag in river bed-sediments whilst we are

less concerned about the river water column. The aquatic discharge of Ag has been noted for many

years as different industrial processes have waxed and waned, most notably with the photographic

industry. Despite its antimicrobial properties this discharge has not appeared to jeopordise the

biological performance of sewage treatment plants.) To in part allay the concerns regarding

sediments and the aquatic environment we would recommend utilising an existing continuous 7

year fish/otter tissue archive to examine whether silver is increasing or declining in British aquatic

wildlife in proportion to the apparent growth in the nanosilver industry. Another recommendation

would be to directly measure Ag in agricultural soils where sludge is applied.

1 Introduction

The overall rationale for these studies was to test the current exposure of the British natural

environment to silver. Our previous study for Defra had established that concentrations likely to

reach the water column do not give cause for concern (Johnson et al., 2014). However, it is less

clear to what extent soil and river sediment are being contaminated with silver to levels which

might be harmful. Such studies could provide valuable context and enable us to judge to what

extent the environment could accommodate further discharge of nanosilver products in the UK.

Thus, if environmental concentrations of Ag were already dangerously high, new Ag applications

would give rise to greater concern. Thus, the tests we proposed provide preliminary information on

the current exposure of the soil and sediment to anthropogenic silver.

2 Sewage sludge sampling and silver analysis

2.1 Materials and methods

2.1.1 Sampling sites:

Anaerobically treated and partially dewatered sewage sludge known as “cake” was collected from

eight sewage treatment works and one dedicated sludge processing facility across England. In all

cases the sludge included biosolids from other sewage treatment works in the area, which has been

transported there for treatment normally by road tanker, but in the case of the dedicated sludge


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treatment facility also by pipeline (Belshaw, 2000). Relationships with the appropriate water

Utilities had been initiated in a previous Defra funded study on nanosilver in sewage effluent and

water (Johnson, 2011; Johnson et al., 2014). The intention had been to find a representative

spread both geographically and of treatment types for the UK. Following discussions with the

Water Utilities the final selection of sewage treatment plants (STPs) was made. Naturally, the

greater the number of STPs sampled the better might our understanding of the UK situation

become, but research resources are never unlimited and it was considered this sample should give

us a reasonable insight of the general picture. It was not intended to find worst case sludges which

might be associated with particular local industries. Indeed in this modern era it is not necessarily

clear which industries might be the greatest polluters (Amneklev et al., 2014)?

All samples of “cake” were of the final product which is ready to be used in agriculture, or

landscaping etc. In many cases the product was low enough in pathogens and heavy metals to be

“Class A”, i.e. suitable for any use, including application to agricultural land used for food

production, but often it was not (yet) marketed as such due some checks or certificates still being

outstanding (personal communication by treatment works staff). Class A sludge has restrictions on

the content of 10 heavy metals but this does not include Ag. Since sludge treatment is a very slow

process (often months) and the sludge treated at one site is a mixture from a number of sewage

works, each individual sample is already integrated over space and time to a certain extent. Figure

1 and table 1 give an overview of the sampling sites.

Figure 1 Location of the sludge sampling sites. Sewage works with sludge treatment in blue, dedicated sludge

treatment plant in green


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Table 1 Sites where sludge samples were taken

sampling site date % moisture

sludge cake comments

Oxford 09/10/2012 21.2%

Aylesbury 24/10/2012 16.3%

Minworth 13/11/2012 18.8%

Shell Green 21/11/2012 26.3%

Dedicated sludge processing plant, treating

sludge from many works in the whole of the

North of England. Sludge arrives by

pipeline. Incoming sludge is about 3% solids

and then gets concentrated to 27%

Blackburn 22/11/2012 29.2%

Ellesmere Port 21/11/2012 23.4%

Fazakerley 27/11/2012 29.1%

This sewage works receives a large

proportion of industrial influent which is

pre-treated in a separate stream before it is

added to the predominantly domestic

influent

Gt. Billing

(Northampton) 13/12/2012 26.6%

Anglian Waters largest sludge treatment

centre. Uses enhanced enzymic hydrolysis

followed by pasteurisation and finally

anaerobic digestion.

Produces Class A sludge but not yet

certified for class A.

Some digested sludge is mixed with garden

waste for co-composting

More than 90% of Anglian Water sludge is

recycled to agriculture (Power et al., 2010)

Cotton Valley (Milton

Keynes) 13/12/2012 30.4%

employs the Cambi process where sludge is

heated at high temperature and pressure

before anaerobic digestion to remove

pathogens and improve the digestion

efficiency

2.1.2 Sampling and dry weight

Due to budget constraints each site was only sampled and measured once, but care was taken to

make this sample as representative as possible:

Sludge cake was collected in the autumn of 2012 in a plastic bucket using a spade and taking

samples from a few different areas of the pile. The bucket was closed with an airtight lid and kept

cool until processing in the laboratory.


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Using a clean spoon, a subsample was taken from different parts of the bucket and placed in

portions on a new sheet of aluminium foil and wet weight determined. The sample was then dried

at 105°C before being weighed again to determine the water content.

Dried sample was then ground in an agate ball mill before being digested and analysed as described

below.

2.1.3 Certified Reference materials

Two certified reference materials were included with each batch: LGC 6181 (dried digested sewage

sludge, LGC, Teddington, UK), which has a silver content of 55 µg/g (close to maximum value found

in the sludge samples below) and NCS 73034 (NCS Testing Technology, China National Analysis

Centre for Iron and Steel, Beijing, China) which has a certified silver content of only 0.13 µg/g.

2.1.4 Sample digestion

1.5 g of dried sample or certified reference material was placed in a 100 ml Kjeldahl digestion tube,

blanks of acid washed sand (HNO3 and oleic acid) were also included with each batch. 10.5 ml

concentrated hydrochloric acid and 3.5 ml concentrated nitric acid (both Aristar grade, VWR, UK)

were added to the Kjeldahl tube, closed off with a glass bubble or marble and left to start the

reaction at room temperature overnight in a fume hood. Then the tubes were placed in a digestion

block and heated to 140C for 2.5 hours.

Once cooled down, the sample was filtered through copper nitrate treated filter paper (Whatman

no 540, soaked for 30 min in 0.1M copper nitrate solution and then washed with ultrapure water),

ensuring complete transfer by rinsing the digestion tubes and closure marbles several times with

0.5M nitric acid and adding this to the sample to be filtered and adding more 0.5M nitric acid to the

filter after the sample had passed through. Then the samples were made up to 100 ml with 0.5M

nitric acid and stored in the dark until analysis by ICP-MS.

2.1.5 Silver Measurement by ICPMS

The digests were analysed using a Perkin Elmer Elan DRC II inductively coupled plasma mass

spectrometer (ICPMS) instrument. The ICPMS instrument was calibrated using Ag standards over

the range 0-10 µg/l in 1M HCl matrix (Aristar) and using 115In as internal standard to compensate for

matrix effects and possible drift in instrument sensitivity. The ICPMS instrument detection limit is

0.14 µg/l (mean blank + 3σ for the reagent blank, n=10) which corresponds to a detection limit of

0.01 mg/Kg in the dried sludge. The instrument method has a precision of 1.4 % (coefficient of

variation, nominal Ag concentration 5 µg/l, n = 10). A typical ICPMS calibration is shown in figure 2.

The sludge concentrations were corrected by the average blank value (acid washed sand, Table 2),

which was <5% of the concentrations found in the sludges. Further discussion of the origin behind

this approach for Ag analysis can be found in previous studies for Defra (Johnson, 2011; Johnson et

al., 2014).


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Figure 2 Calibration curve for silver using standards in the range 0-10 µg/l.

2.2 Results

Table 2 and 3 show the measured silver concentrations in the reference materials and sludge

samples respectively. Note that the quantification ion is 107Ag with 109Ag being used to confirm the

results.

Table 2 certified reference materials and blanks. All values in mg/Kg dry weight (dw)

sample

and expected value Ag (107) Ag (109) confirmation ion

sand (blank) 0.074 0.003

-0.045 -0.002

0 0.343 0.346

-0.017 -0.019

0.216 0.226

0.102 0.101

average (SD) of all blanks 0.112 (0.146) 0.109 (0.148)

CRM NCS 73034 0.189 0.195

0.194 0.219

0.13 0.176 0.204

0.225 0.447

0.139 0.335

average (SD) NCS 73034 0.1846 (0.031) 0.280 (0.109)

average recovery 142% 215%

y = 14013x + 381.73 R² = 1

0

20000

40000

60000

80000

100000

120000

140000

160000

0 2 4 6 8 10 12

Inte

nsi

ty, C

PS

Ag, µg/l

ICPMS Calibration for Ag


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sample

and expected value Ag (107) Ag (109) confirmation ion

CRM LGC 6181 53.9 53.4

53.5 53.3

55 54.0 53.8

54.2 52.3

61.4 59.6

average (SD) LGC 6181 55.4 (3.3) 54.5 (2.9)

average recovery 101% 101% a this is close to the quantification limit of the method

Table 3 Silver concentrations in sludge cake from 9 sewage or sludge treatment works

sludge site and date date

Ag (107)

[mg/Kg dry weight]

Ag (109) confirmation ion

[mg/Kg dry weight]

Oxford

09/10/12 3.48 (3.36, 3.59) a 3.60 (3.48, 3.72)

Aylesbury 24/10/12 3.56 3.75

Minworth 13/11/12 7.34 (7.51, 7.17) a 7.46 (7.61, 7.30)

Shell Green 21/11/12 4.62 4.80

Blackburn 22/11/12 2.88 3.04

Ellesmere Port 21/11/12 2.34 2.48

Fazakerley 27/11/12 42.10 41.8

Gt. Billing 13/12/12 3.98 (4.10, 3.85) a 4.10 (3.99, 3.08)

Milton Keynes 13/12/12 2.92 3.08

Low: 10% ile 2.77

Median 3.56

High: 90% ile 14.3 a duplicate digestions, the average and the individual measurements are given

There was a good agreement between the duplicate analyses and between the quantification ion

Ag(107) and the confirmation ion Ag (109).

The recoveries for the high reference material LGC 6181 was close to 100% (average: 101%, range:

97%-112%), but for the low one with 0.13 mg/Kg Ag was close to the quantification limit and the

recoveries were not as good (average: 142% ). However, the silver concentrations in the samples

were a factor 18-300 higher than this low standard and up to a factor of 20 lower than the higher

reference material, so they fell in a region where the analysis was expected to perform well.

The values for Ag in sludge at all the sites except Fazakerley were similar. The Fazakerley value was

6-10 fold higher than for the other STPs and perhaps this was linked to its high industrial input

compared to other plants in the study.


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Figure 3 concentration of Ag in sludge cake from 9 different sites. For Oxford, Minworth and Gt. Billing, two

sludge sub-samples were analysed and the bar gives the average with the error bars indicating the individual values

2.3 Discussion of sludge results

The values found in sludge were 2.77 mg/kg dw (10%ile), 3.56 mg/kg dw (median) and 14.3 mg/kg

dw (90%ile). The only other studies we have managed to find on Ag in sewage sludge was for one

STP in Sweden with 8 mg/kg dw (Ahlberg et al., 2006) and a median of 17 mg/kg from 14 STPs

sampled in Japan in the 1990’s (Kawasaki et al., 1998). The highest value found at one of these

Japanese STPs was 61 mg/kg Ag. Thus, this project has increased the world published knowledge

on Ag in sludge by almost 40%! A previous predictive exercise suggested values of 1.3 to 4.4 mg/kg

nanosilver might be found in European sludge (Gottschalk et al., 2009), which are not dissimilar to

our values for total silver. We can compare these Ag sludge measurements with estimates from the

previous Defra funded study (Johnson et al., 2011) where silver was measured in the influent and

effluent of 9 STPs. In this case the amount of silver found on the particulates (>0.45 µm) in the

influent might simulate what could be found in sewage sludge (although they only represent a brief

snapshot). These particulate values from the previous study are in a comparable range to the

sludge measurements in this study (Figure ). As might be expected from their greater spatial and

temporal integration, there was less variability in the sludge cake samples than for the previous

0.45 m influent filtrate measurements, with only Fazakerley standing out for its high silver

content. Overall, the range of values found here fall within that reported in the existing but limited

world literature.

3.5 3.6

7.3

4.6

2.9 2.3

4.0

2.9

0

2

4

6

8

10

Oxford09/10/12

Aylesbury24/10/12

Minworth13/11/12

Shell Green21/11/12

Blackburn22/11/12

EllesmerePort

21/11/12

Fazakerley27/11/12

Gt. Billing13/12/12

MiltonKeynes

13/12/2012

Ag

[ m

g/kg

dw

] 42


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Figure 4 Silver concentrations measured in sludge cake from this study (blue) compared with those previously

measured in suspended solids in the sewage influent (red).

2.4 Estimating silver concentrations building up in soils

We reported previously in the first task of this study (Williams and Johnson, 2012) estimates for

silver build up in soils. These previous calculations were based on the amount of silver (solid or

colloidal) found in the influent together with estimated removal rates from the liquid phase into the

sludge and 62 % of sludge solids being applied to agricultural land, which was based on a report

from 2006 (WaterUK, 2006). More recent data (Defra, 2012) shows that this proportion has

increased to around 80% since then, with sludge use as fertilizer being promoted as the most cost-

effective and environmental disposal route (Figure 5).

Using the median concentration of 3.56 mg/Kg or g/t (Table ) and the latest figures for sludge

disposal to land from Figure 5, a median silver application to land of 1,118,159 t sludge x 3.56 g/t Ag

= 3,980,646 g or 3.98 t Ag applied to land every year can be calculated. Similarly, using our 10% and

90% Ag sludge measurements, the lowest and highest estimates are 3.10 and 16.0 t Ag to land/yr

respectively. This compares well with the previous predictions of a median 5-6 t/yr Ag to land and

10% and 90% Ag sludge of 1 and 39 t/yr made previously in this project. It is important to recall this

is total silver, the latest prediction for total annual nanosilver loss to the sewers across Europe is

8.8 t/yr (Sun et al., 2014)

3.5

3.6

7.3

4.6

2.9

2.3

42

4.0

2.9

0.6

9

8.3

37

11

.4

0.9

2

0.8

2

39

5.1

1.3

6.0

0

5

10

15

20

25

30

35

40

45

Ag

mg/

Kg

dw

sludge cake, 2013 influent suspended particles >0.45 µm, 2010


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Figure 5 re-use and disposal routes for UK sewage sludge (Defra, 2012). Values in tonnes dry solid per year.

Table 4 Predicted loads of total Silver particles applied to agricultural soils in the UK based on sludge measurements.

Ag in sludge

cake [mg/Kg

dw† ]

Ag mass to

agricultural land

from sludge cake

estimates† [t/yr]

Ag µg/Kg dw

accumulated in

soil per year *

10%ile 2.77 3.10 8.8

median 3.56 3.98 11.3

90%ile 14.3 16.0 61.5

* assuming that sludge is incorporated into the top 20 cm, the bulk dry density is 1.3 t/m3 and the maximum

allowed sludge is applied every year

Good agricultural practice advises limiting total N applications to 250 kg/ha/year N. As sludge is

considered to contain a minimum of 3% N by dry weight (Hogan et al., 2001), up to 8.3 tonnes

dw/ha sludge may be applied. As we predict 2.77, 3.56 or 14.3 mg/Kg dw Ag for the 10%ile, median,

and 90%ile estimates respectively (Table 4) that would result in an inadvertent application of silver

of up to 23.0, 29.5, 123 g/ha or (2.30, 2.95, 12.3 mg/m2 Ag) if this agricultural practice was

followed. The typical plough depth is 20 cm and a typical soil dry weight bulk density is 1.3 t/m3

(Hallett, 2008), so 1 m2 ploughed to 20 cm gives a soil volume of 0.2 m3, or 260 kg by weight of soil.

440,137 t 44%

1,241,639 t 81%

1,118,159 t 79%

0

200,000

400,000

600,000

800,000

1,000,000

1,200,000

1,400,000

1,600,000

1,800,000

1992 2008 2010

Other disposal

Incineration

Landfill

Other re-use

Soil & agriculture

Ships to sea

Pipelines to sea


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Thus, for the median value of silver in sludge we calculate 2,950 µg Ag/260 kg soil = 11.3 µg/Kg

concentration of Ag in soil (Table 4). This value can be used to calculate loads if sludge is repeatedly

applied to the same field. Background silver measurements for topsoil are recorded in the

European geochemical baseline database (FOREGS, Salminen et al., 2005). For the UK these values

range from 0.14 mg/kg to 1.2 mg/kg with a median of 0.3 mg/Kg. Thus sewage sludge additions

would only add up to 4% more silver to the existing UK soil background assuming the FOREGS data

is representative (using median silver concentrations for soil and sewage sludge (Table 4) and

maximum annual sludge application for the calculations). A study using 0.6 m deep lysimeters

applied with sludge and then irrigated found only very small quantities of Ag leached through the

soil over the course of a year, indeed Ag mobility was in the same bracket as Cd and Pb (Ahlberg et

al., 2006). This limited mobility was also found in a study of some American soils (Yang et al.,

2014).

There are still only few studies on the ecotoxicity of ionic and nanosilver in soil. The recently

completed NanoFATE project combined literature and its own lab studies to generate species

sensitivity distributions for Ag and nano-Ag. The HC5 (level at which the most sensitive 5% of

species would be affected) for the ion form of Ag in soil was 6.2 mg/kg and from this a PNEC of 1.24

mg/kg ionic Ag was derived. The lowest effect concentration where 50% of the individuals died

(EC50) in soil studies in the NanoFATE project was reported as 4.8 mg/Kg for nanosilver. Thus,

according to the calculations in this study, a single sludge application to soil would introduce Ag at a

level 100 times below the suggested PNEC for ionic Ag of 1.24 mg/kg. But repeated sludge

application to the same soil (over ten times) might start to reach levels of concern. The toxicity of

Ag has been very closely linked to concentrations of the free silver ion (Ratte, 1999). However, in

the real world Ag entering STPs is considered to almost entirely react with sulphide to form stable

precipitates (Choi et al., 2009; Kim et al., 2010; Kaegi et al., 2013). These have the effect of

reducing the release of the toxic ion from particulates and nanoparticulates and so reduce the

toxicity (Levard et al., 2011; Levard et al., 2013). This implies that current soil toxicity studies may

overestimate the toxicity of Ag entering the soil from sludge.

There is still much research to be done on the fate and behaviour of nanosilver in soils. However,

the addition of silver to soil from sludge does not appear to be an immediate concern. We are not

clear to what extent repeat sludge applications are normal agricultural practice in the UK but this

may require further consideration.


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3 Presence of silver in river bed-sediment

3.1 Materials and methods

3.1.1 Sampling sites:

It had been planned to sample from nine separate river sites rather than seven. Unfortunately,

developing a successful freeze-coring technique for deep river sampling took time. Following the

seventh sample prolonged flooding took place which precluded further sampling. Nevertheless a

good series of contrasting locations were sampled. Sampling took place in the Thames catchment

which includes deeply rural sites with little industry and modest human populations such as at the

Kennet and Ock and near to urban sites with industry such as Littlemore Brook (Oxford) and The

Cut (Bracknell).

Figure 6 Overview of the sampling sites in the Thames catchment

Littlemore Brook x2

Thames @Wallingford x2

Kennet

Ock

The Cut


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Figure 7 Detail of the Littlemore Brook sites

Figure 8 Detail of the Wallingford sites

3.1.2 The freeze-coring method

The intention of using freeze-coring to sample river bed-sediments was to retrieve intact and

representative segments of the river bed. This gives the opportunity of preserving the historic

strata of the river bed, if they actually exists. The first stage is the actual coring followed by

sectioning the core into a series of segments of known depth.

A pointed core made from 28 mm diameter copper pipe was hammered into the sediment to a

depth of 30-40 cm. This hollow core was filled with about 80 ml methylated spirits (to aid heat

transfer) and dry ice (solid CO2). The core was kept topped up with dry ice for up to 1 hour to

ensure a sufficient thickness of sediment to freeze to it. Then the core with attached sediment was

pulled out using a winch, packaged in a plastic bag and returned to the laboratory packaged in

industrial area

Thames


- 15 -

more dry ice. The corer with attached sediments was stored in a freezer until the sample could be

processed. The sediment core was removed from the copper pipe by filling the pipe with hot water

until it was possible to slide the sediment core off and return it to the freezer. Two to five cores

were taken at each site, but due to the limited resources allocated to the project, only one could be

analysed (two at Littlemore Brook because on the first attempt the top layer did not freeze).

With this technique, sediment samples can only be taken from water depths of less than about 1m.

Therefore samples could not be taken right in the middle of most streams, but with the exception

of the two sites on the Thames itself much of the width of the streams was suitable.

The first of the Thames samples (ds. Bradfords Brook, right bank) was divided into sections by

cutting the frozen sediment core with a handsaw, all other samples were left to defrost over night

on a flat surface and then divided into sections using a kitchen knife. Rather than using fixed

thicknesses for the depth sections, cuts were made where the structure of the sediment changed

visually. Additional cuts were made if no visual changes occurred over distances of about 10 cm or

more.

The sediment sections were dried in an oven at about 90°C, and then they were broken up by hand

and sieved to 2mm mesh size, before grinding and digesting them as described above for the

sewage sludge.


- 16 -

Figure 9 schematic setup of freeze corer

EtOH + dry ice

insulation

28 mm copper pipe


- 17 -

a) b)

Figure 10 a) adding the dry ice b) retrieving the sediment core

a) b)

c) d)

Figure 11 a) corer with pipe insulation marking the required depth, b) frozen core, c) defrosted core, e) slicing

3.1.3 Silver analysis

The digestion of river bed sediments followed the procedure used for sewage sludge and already

described in section 2.1.4. However, the final measurement of Ag concentration differed slightly


- 18 -

with the digests being analysed using a Perkin Elmer Nexion 300D inductively coupled plasma mass

spectrometer (ICPMS) instrument. The ICPMS instrument was calibrated using Ag standards in 1M

HCl matrix (Aristar) over the range 0-10 µg/l and using 115In as internal standard to compensate for

matrix effects and possible drift in instrument sensitivity. The ICPMS instrument detection limit is

0.1 µg/l (mean blank + 3σ for the reagent blank, n=10), which is equivalent to 0.006 mg/Kg in the

sediments and the instrument method has a precision of 1.4 % (coefficient of variation, nominal Ag

concentration 5 µg/l, n = 10). A certified reference material LGC 6181 (dried digested sewage

sludge, LGC, Teddington, UK, 55 mg/Kg silver) was run after every 8 samples and gave average

recoveries of 94%.

3.1.4 Cs-137 analysis of sediment samples

Dry sediment samples were accurately weighed into either 15 ml clear Petri dishes or 150 ml plastic

containers depending upon the sample volume available. All samples were analysed as received

(sieved to 2mm, dried and then finely ground) for four days and were decay corrected to the day of

sampling. Samples were analysed using hyper-pure germanium detectors coupled to a

computerised analytical system. The detectors are calibrated for efficiency using a mixed

radionuclide standard which covers an energy range of approximately 59 -1850 keV

(http://www.npl.co.uk/upload/pdf/mixed_nuclide_standard_solutions.pdf). The resultant spectra

were analysed using Canberra Apex-Gamma software and the activity concentrations (and 2σ

counting errors) of 137Cs estimated. The method is accredited by the UK Accreditation Service

(methods SOP4504 and SOP4505; see http://www.ukas.org/testing/schedules/Actual/

2506TestingSingle.pdf for full schedule).

3.1.5 Loss on ignition

To estimate the organic carbon contents of the sediments, the organic carbon was burnt off by

firing the dried sediments (dried at 105°C) at 550°C and the determining the weight loss (SOP 3101

issue 6).

http://www.npl.co.uk/upload/pdf/‌mixed_nuclide_standard_solutions.pdf

http://www.ukas.org/testing/schedules/Actual/2506TestingSingle.pdf

http://www.ukas.org/testing/schedules/Actual/2506TestingSingle.pdf


- 19 -

3.2 Results for river bed-sediments

3.2.1 Silver contamination of river bed-sediment samples

The average silver recovery from the certified reference material was 94%. With the exception of

the site with the highest silver concentrations, all sediment samples were within the range of silver

concentrations previously recorded for soils (topsoils 0.14-1.2 mg/kg and subsoils 0.02-2.1 mg/kg)

in the UK (Salminen et al., 2005). The site with the highest Ag content (5 mg/kg) was that

downstream of the major sewage treatment plant for Oxford described here as Littlemore Brook

downstream (see Figure 7 for location, Figure 12 and Figure 15 for silver content). This, together

with the peaks at Thames at Bradford Brook and the Cut (Figure 20), suggests the major source for

Ag in the Thames catchment is from the human population/industry. The highest Ag

concentrations at Littlemore Brook downstream were found between 30 and 45 cm depth, in a

location similar to the detectable Cs-137. It is, therefore, tempting to suggest that the highest Ag

discharge we found relates to a historic peak in losses in the 1950s-1960s (see Section 3.2.2),

perhaps from the photographic industry or other industrial processes. The worldwide silver use in

photography peaked in 1999 (http://goldnews.bullionvault.com/silver-bullion-photographic-

demand-062120133), but due to improved recovery of this expensive material in the film

processing plants as well as improved sewage treatment it is likely that losses to the environment

peaked significantly earlier. This deep high Ag fraction is unique to this Oxford sewage discharge

site. The sites with the highest Ag concentrations in the upper (presumably most recent) layers (0.7

mg/kg) are in the more urbanised locations or closest to sewage discharge points at Littlemore

Brook (immediately downstream of where the 350 m long sewage effluent channel enters the

brook), Thames near Bradford’s Brook (which passes the small industrial area at the outskirts of

Wallingford) and The Cut (which is heavily influenced by the urban area of Bracknell as well as being

only 1.5 km downstream of the effluent channel from Bracknell sewage works) (Figure 13). The

difference in Ag content between these locations near urban centres and the control sites on the

Ock and Kennet was 10-20-fold (Figure 12). We are not aware of data for background silver

sediment concentrations for the UK, but the levels measured are similar to those previously found

in soils (FOREGS project, Salminen et al., 2005).


- 20 -

Figure 12 Average (blue bar), standard deviation (whisker) and maximum silver concentrations at the different sites

Figure 13 Silver concentration in the slice that includes 8 cm depth, to help compare different sites at a given depth

Overall, the problem of sediment disturbance makes it difficult to say whether Ag concentrations in

river sediment are increasing or decreasing, but we can say that the major source of Ag today in

the Thames is likely to be human activity. Furthermore, at most sites the silver concentrations were

strongly correlated with the organic carbon content as estimated by loss on ignition (Figures 14 –

20). Therefore, the highest silver concentrations within a river stretch are likely to be found where

organic-rich sediments have settled. However, between rivers, normalising the silver concentration

to organic carbon content does not greatly reduce the variability (Figures 21 and 22).

LittlemoreBrook upstr.September

andOctober

LittlemoreBrook

downstr.September

andOctober

ThamesWallingfordbeach, ds of

W.fordBridge

ThamesWallingford

dsBradford's

Brook

Ock atCaldecott

Kennet atNorth Croft

The Cut @Bracknell

silver max 0.756 5.290 0.041 1.44 0.038 0.107 0.789

silver average 0.199 1.189 0.036 0.73 0.031 0.057 0.308

0

1

2

3

4

5

6mg/Kg

LittlemoreBrookupstr.

Septemberand

October

LittlemoreBrook

downstr.September

andOctober

ThamesWallingford

beach, dsof W.ford

Bridge

ThamesWallingford

dsBradford's

Brook

Ock atCaldecott

Kennet atNorth Croft

The Cut @Bracknell

silver at ca. 8 cm 0.756 0.117 0.037 0.65 0.030 0.107 0.789

00.10.20.30.40.50.60.70.80.9

mg/

Kg


- 21 -

Figure 14 Ag and organic material in Littlemore Brook sediment upstream of the major sewage treatment plant

The Littlemore Brook upstream site whilst being upstream of the sewage discharge point is still

within an urbanised and industrial sub-catchment of the Thames. Perhaps some other domestic or

industrial waste enters the stream and sediments. The Ag and sediment organic content (LOI) were

closely related (Figure 14) with high (15%) organic content near the surface. This relationship might

indicate the Ag and organic material have become bound together or reflect a co-relationship of

fine sediments.

Figure 15 Ag and organic material in Littlemore Brook downstream of the major sewage treatment plant

The downstream site at Littlemore Brook near the sewage effluent pipe for Oxford appeared to

have modest Ag in the top layers but very high levels of up to 5 mg/kg in a deeper layer (Figure 15).

Possibly this Ag peak reflects heavy use with little recovery from the waste water in an earlier

industrial period which coincided roughly with the 1950s-60s nuclear fallout period (Table 5). For

0% 5% 10% 15% 20% 25%

-60

-50

-40

-30

-20

-10

0

-60

-50

-40

-30

-20

-10

0

0 0.2 0.4 0.6 0.8 1 1.2 1.4 1.6

Loss on ignition d

ep

th [

cm]

silver [mg/Kg]

Littlemore Brook upstream

silver Oct 2013 silver Sept 2013 LOI Oct 2013 LOI Sept 2013

0% 5% 10% 15% 20% 25%

-60

-50

-40

-30

-20

-10

0

-60

-50

-40

-30

-20

-10

0

0 1 2 3 4 5 6

Loss on ignition

de

pth

[cm

]

silver [mg/Kg]

Littlemore Brook downstream

silver Oct 2013 silver Sept 2013 LOI Oct 2013 LOI Sept 2013


- 22 -

example silver use in photography peaked in 1999, but already before technologies to minimise the

losses of the expensive raw material silver were developed, such that the peak in release was likely

to be significantly earlier.

Figure 16 River Thames main stream at Wallingford

The Wallingford Bridge site is on the main stem of the middle of the Thames and shows very low Ag

content of 0.04 mg/kg DW and average organic content (5%) shown in Figure 16. There was a

general similarity between the Ag and organic profile in the sediments in the Thames here.

Figure 17 River Thames main stream at Wallingford downstream of Bradford’s Brook.

The Thames site close to Bradford’s Brook shows Ag levels almost 30 X higher than the other

Wallingford site (Figure 17). Bradford’s Brook passes the small industrial area on the outskirts of

Wallingford before flowing into the Thames. The upper half of this core was characterized by pieces

0% 5% 10% 15% 20% 25%

-40

-35

-30

-25

-20

-15

-10

-5

0

-40

-35

-30

-25

-20

-15

-10

-5

0

0.00 0.02 0.04 0.06 0.08 0.10 0.12 0.14 0.16

Loss on ignition

de

pth

[cm

]

silver [mg/Kg]

Thames Wallingford beach, ds of W.ford Bridge

silver, mg/Kg dw LOI, %

0% 5% 10% 15% 20% 25%

-30

-25

-20

-15

-10

-5

0

-30

-25

-20

-15

-10

-5

0

0 0.2 0.4 0.6 0.8 1 1.2 1.4 1.6

Loss on ignition

de

pth

[cm

]

silver mg/Kg

Thames Wallingford ds Bradford's Brook



- 23 -

of wood presumably dropped from nearby trees leading to the highest organic carbon content (loss

on ignition) of any of the analysed samples. There was a strong correlation between organic carbon

and silver content.

Figure 18 River Ock a rural tributary of the Thames upstream of Oxford

The site at the River Ock is a rural area with little sewage input and this appears to be reflected in

the low Ag content of 0.04 mg/kg Ag similar to the Wallingford Bridge site (Figure 18). There was a

general similarity been the Ag and organic profile in the sediments in the Ock.

Figure 19 River Kennet at Newbury, a rural stretch of a tributary of the Thames (Marlborough, Newbury then joins at

Reading)

The site on the Kennet is also a rural location which has some peaks of Ag at 0.1 mg/kg but mainly

below 0.04 mg/kg Ag (Figure 19). Unusually, at this site the Ag and organic (LOI) profiles do not

correspond. Possibly there is a historic presence of Ag at the 35 cm depth?

0% 5% 10% 15% 20% 25%

-40

-35

-30

-25

-20

-15

-10

-5

0

-40

-35

-30

-25

-20

-15

-10

-5

0

0 0.02 0.04 0.06 0.08 0.1 0.12 0.14 0.16

Loss on ignition

de

pth

silver mg/Kg

Ock at Caldecott


0% 5% 10% 15% 20% 25%

-40

-35

-30

-25

-20

-15

-10

-5

0

-40

-35

-30

-25

-20

-15

-10

-5

0

0 0.02 0.04 0.06 0.08 0.1 0.12 0.14 0.16

Loss on ignition

de

pth

silver mg/Kg

Kennet at North Croft



- 24 -

Figure 20 The Cut river, a very urbanised tributary of the Thames receiving waste from Bracknell

The site on the Cut River is an urbanised tributary receiving waste from Bracknell and so it has Ag

levels between 0.2 and 0.9 mg/kg DW suggesting high exposure from human activity (Figure 20).

There is a general similarity between the Ag and organic profile in the sediments.

Figure 21 Silver concentrations at all sites on a logarithmic scale

0% 5% 10% 15% 20% 25%

-40

-35

-30

-25

-20

-15

-10

-5

0

-40

-35

-30

-25

-20

-15

-10

-5

0

0 0.2 0.4 0.6 0.8 1 1.2 1.4 1.6

Loss on ignition d

ep

th [

cm]

silver [mg/Kg]

The Cut @ Bracknell


-70

-60

-50

-40

-30

-20

-10

0

0.01 0.1 1 10 100

de

pth

silver mg/Kg

All sites , logarithmic scale

Thames Wallingford ds Bradford's Brook

Thames W'ford beach, ds of W.ford Bridge

Ock at Caldecott

Littlemore Brook upstr. September

Littlemore Brook upstream October

Littlemore Brook downstr. September

Littlemore Brook downstream October


The Cut at Bracknell


- 25 -

Figure 22 Organic carbon normalised silver concentration at all sites on a logarithmic scale

3.2.2 Cs-137 in sediment samples

Radioisotopes from weapons fallout have been used previously to identify sediments laid down in

the mid 1950's to the 1960's (Bothner et al., 1998). The UK also received Chernobyl fallout (1986)

so a peak associated with 1986 may also occur. An attempt to consider the ages of the sediment

cores using the Cs-137 activity data was made in a manner consistent with the method described in

Jeter (2000). The deepest occurrence of detectable Cs-137 in sediment profiles generally marks the

year 1954 (i.e., start of atmospheric testing of nuclear bombs), while peak activities correspond to

sediments from 1963 (Simpson et al., 1976). Williams (1995) used Cs137 for age determination,

saying that the peak equals 1954, so above and below is before and after 1954. The levels reported

in Table 5 are typical for weapons fallout and a Chernobyl deposition in soils in the area (see CIG

view services linked from https://wiki.ceh.ac.uk/x/L4FsD and https://wiki.ceh.ac.uk/x/G4FsD). It is

likely that weapons fallout dominates compared to the Chernobyl fallout in this area, the latter

contributing a maximum of around 30%. Due to the relatively small sample sizes available, Cs-137

levels were mostly near or below the detection limit of the method. Not enough sections could

reliably be quantified to be able to determine whether the 1954 or 1986 peaks exist or whether

older and newer sediments have been mixed. This contrasts with apparently more stable locations

-70

-60

-50

-40

-30

-20

-10

0

0.1 1 10 100 1000

de

pth

silver mg/(Kg organic carbon)

All sites , organic carbon normalised, logarithmic scale Thames Wallingford ds Bradford's Brook

Thames W'ford beach, ds of W.ford Bridge

Ock at Caldecott

Littlemore Brook upstr. September


Littlemore Brook downstr. September

Littlemore Brook downstream October



https://wiki.ceh.ac.uk/x/L4FsD

https://wiki.ceh.ac.uk/x/G4FsD


- 26 -

such as Boston Harbour (USA) where a distinct peak at 20-40 cm depth was found (Bothner et al.,

1998).

Table 5 Distribution of Cs137 with bed-sediment depth

Site depth (cm)

number sections

section 1 (Bq/kg DW)

2 (Bq/kg DW)

3 (Bq/kg DW)

4 (Bq/kg DW)

5 (Bq/kg DW)

6 (Bq/kg DW)

7 (Bq/kg DW)

8 (Bq/kg DW)

Littlemore Brook upstream

September 15-63 11 <1.80


0-26 4 3.11 <1.42

Littlemore Brook downstream September

25-63 7 <0.89 <1.89 <1.33 <0.82 <1.89 1.13 <0.76

Littlemore Brook downstream

October 0-23 4 <1.85

<1.29

Thames W.ford beach, ds. of W.ford Bridge

0-39 11

2.12

Thames Wallingford, ds. of Bradford’s Brook

a 0-30 6 3.97 3.18 2.75 <4.48 <3.35 <1.31

Ock at Caldecott 0-37 6 <0.68 <1.48

Kennet at North Croft (Newbury)

0-38 8 <1.63 <1.97 1.87 <0.77 <2.58 <2.20 <1.02 <1.06


0-42 7 <1.36 2.45 <2.08 1.51 2.19 <2.24 <1.32

a Values from this site are not UKAS accredited. empty fields: this section does not exist or was not analysed

We expect sediments to arrive in rivers from soil erosion, bank erosion, sewage particulates and

indigenous organic matter particles. The rate of sediment deposition will be related to the particle

size and the water velocity. However, unlike for lakes, significant disturbance can occur due to

violent high flow events picking up and moving sediments in rivers and streams and in sometimes,

especially in the larger rivers, due to dredging. Thus, we would expect Cs-137 to have been

deposited in British soil from the late 1950s peaking in 1963 (and subsequently in 1986). Bed-

sediment disturbance may be more extreme in one location compared to another. In the Thames

at Wallingford Cs-137 is detected in the top 10 cm but not below this. At Littlemore Brook

upstream Cs-137 was found in the top layer. At Littlemore Brook downstream of the sewage

discharge point Cs-137 was found only in a deep layer at about 50 cm. The Cut at Bracknell has Cs-

137 at 5-10 cm and also at 15-25 cm. However, it should be noted that all concentrations are close

to the limits of detection and since the limits of detection depend on the amount of sample


- 27 -

available, non-detects have not necessarily lower concentrations than those sections where Cs-137

was measurable.

4 Overall conclusions

We can now compare exposure to Ag from this study in sludge, soil and river bed-sediments

(Table 6):

Table 6 Comparison of Ag detected in sludge, predicted in soil and found in river bed-sediments (all units as mg/kg

DW)

Measured

median total Ag

in sludge cake

Predicted annual

total Ag content

in soil from

sludge

application

Median total

Ag content in

British soils

(background)

FOREGS

Range of total

Ag in less human

exposed river

bed-sediments

Range of total

Ag in most

human exposed

river bed-

sediments

PNEC for ionic

Ag for soil

organismsa

(EU Nanofate

project)

3.56 0.011 0.3 0.02-0.04 1-5 1.24

a a PNEC for sediment dwelling organisms has not yet been established

This suggests that river bed-sediments could well have higher Ag exposures than soil, with some

bed-sediments having Ag at levels similar to sewage sludge! This indicates that exposure of river

benthic organisms to Ag might in some locations be greater than other wildlife in the UK. But what

proportion might nano Ag be in this total Ag discharge from sewage plants? In our previous study

carried out for Defra colloidal Ag containing the nano Ag fraction was around 10% of total discharge

(Johnson et al., 2014) and in a study carried out on German STPs this was around 5% (Li et al.,

2013). Thus, nano Ag would appear to be only a small part of total Ag discharge to the

environment so far. Although more coring would be needed for corroboration, it appears the

situation for the aquatic environment was worse in terms of Ag exposure in the 1950s and 60s if

the Oxford Littlemore Brook situation is a guide. Looking at the Boston Harbour (USA) example Ag

in the upper sediments in 1977 was over double what was found in 1993 which also suggests a

downward trend (Bothner et al., 1998). Some authors have argued that more Ag may have been

used in the past (Rozan et al., 1994; Purcell and Peters, 1998; Eckelman and Graedel, 2007) but

even so it is difficult for scientists to keep track of the diversity of Ag applications in industry and

our daily lives with some applications such as wet-chemistry photography disappearing fast, while

other uses such as in electronics and applications of silver nanoparticles are increasing at a similarly

fast rate (Amneklev et al., 2014; The Silver Institute, 2014). The Silver Institute argues that silver is a


- 28 -

better investment than for example gold, precisely because it has so many industrial uses, for which

there is mostly no good alternative, and large price rises over the last decade seem to prove them

right{The Silver Institute, 2014 #391}.


- 29 -

5 Recommendations

5.1 The aquatic environment

We have identified Ag contamination in river bed-sediments associated with domestic or industrial

human activity. Compared to what we think is happening in soils, river bed-sediments appear to be

more exposed. Thus, Ag could be entering the river food web. But we cannot say clearly, due to

the apparently disturbed nature of river bed-sediments, whether silver contamination is going up

or down nor do we have the evidence of whether transfer of silver into river food webs is

significant. Such information on time trends and trophic transfer would be very helpful given the

concern that the Ag nanoparticle industry is a growth area in the UK economy. In a review of Ag

toxicity, bioconcentration factors (BCFs) and trophic transfer, the highest Ag values per body weight

(10,000 BCF) were associated with algae (Ratte, 1999). The review found scientists in laboratory

studies had reported lower BCFs for crustaceans, insects (0.5-1000 BCF but only 0.7-2 with

sediments) and molluscs (8-1000 BCF). For omnivorous fish such as carp a BCF above 1000 was

reported in a study with Ag and sediments. Thus, although not as bioaccumulative as some other

metals and persistent organic pollutants, laboratory studies have shown that transfer from

sediments to aquatic wildlife such as fish is possible. Given the apparently modest sediment BCF

for macroinvertebrates and the very large quantity of biomass required for Ag analysis, measuring

the metal in an omnivorous fish would appear to be the most practical.

At CEH we have collected and frozen fish tissue from wild roach from locations on the Thames,

Nene, Welland, Anker and Glen Rivers, annually since 2007 (http://www.ceh.ac.uk/sci_programmes

/water/nationalfishtissuearchive.html). The sampling sites span both very urbanised and rural

locations. These fish tissue samples have been used successfully to assess Hg contamination

(Jürgens et al., 2013). It might be expected that the period of 2007 to 2014 would span the growth

in the nanosilver products in the UK. CEH’s Predatory Bird Monitoring Scheme (PBMS-

http://pbms.ceh.ac.uk), through its existing links with the Cardiff University Otter Project (CUOP -

http://www.otterproject.cf.ac.uk/), also offers the opportunity to examine trophic transfer of silver.

It would be possible to examine if there has been a time-related increase in liver Ag concentrations

in a selected sample of otters according to location, thereby providing complimentary analysis to

that done on fish. Therefore, it would be possible to analyse a spatially and temporally stratified

http://www.ceh.ac.uk/sci_programmes/water/nationalfishtissuearchive.html

http://www.ceh.ac.uk/sci_programmes/water/nationalfishtissuearchive.html


- 30 -

sample to determine whether otter liver Ag concentrations increase with predicted silver inputs

into river systems from sewage effluent.

Thus, an opportunity exists for retrospective Ag analysis in the tissue of biota from exposed and

rural UK river locations to examine whether Ag concentrations are present and rising in wildlife and

whether these are linked in a predictable manner to inputs. As it is wildlife that ultimately we wish

to protect, this information would be doubly useful.

The outcomes of such analysis, whilst not focused on whether wildlife are suffering from Ag

toxicity, would still have policy relevance as follows:

Possible result 1): Fish and/or otter exposure to Ag is not related to human discharge

This would not support concerns that Ag discharged from the human community, thanks to the

growth in nanosilver applications, is a growing threat to biota, so less additional research needed.

Possible result 2): Fish and/or otter exposure to Ag has not increased over the past 7 years

This would not support concerns that Ag, thanks to the growth in nanosilver applications, is in

reality a growing threat to biota, so less additional research needed.

Possible result 3): Fish and/or otter exposure to Ag has increased over the past 7 years at sites

associated with human activity. This would support concerns that Ag, thanks to the growth in

nanosilver applications, is in reality a growing threat to biota, so more research would be needed.

5.2 The terrestrial environment

This study has developed methods which have reliably found Ag in the complex matrix of sewage

which could be used for soils. We have gone to some lengths to predict Ag in soils where sludge

might be applied as a fertiliser. Given the loads involved and the potential proximity to levels of

concern in soil following repeat applications, it might be prudent to actually measure Ag in soils in

agricultural areas where sludge is used as a soil fertiliser. Such a study could measure at selected

farms; (a) Ag in the sludge product dumped on the soil and waiting to be incorporated; (b) soil that

has not received sludge; (c) soil which has received sludge. Sampling would be depth stratified and

information acquired on the origin and history of the sludge applied.


- 31 -

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