Speciation and bioavailability of zinc in amended sediments

Speciation and bioavailability of zinc in amended

sediments

Aaron G.B. Williamsa, Kirk G. Scheckelb*, Gregory McDermottc, David Gratsonc,Dean Neptunec and James A. Ryand

aEastern Research Group, Inc., 10200 Alliance Road, Ste 190, Cincinnati, OH 45242, USAbU.S. Environmental Protection Agency, 5995 Center Hill Ave, Cincinnati, OH 45224, USAcNeptune and Company, Inc., 8962 Spruce Ridge Rd, Fairfax Station, VA 22039, USAdU.S. Environmental Protection Agency, Retired

*E-mail: [email protected]

ABSTRACT

The speciation and bioavailability of zinc (Zn) in smelter-contaminated sediments were investigated as afunction of phosphate (apatite) and organic amendment loading rate. Zinc species identified in preamend-ment sediment were zinc hydroxide-like phases, sphalerite, and zinc sorbed to an iron oxide via X-rayadsorption near edge structure (XANES) spectroscopy. Four months after adding the amendments to thecontaminated sediment, hopeite, a Zn phosphate mineral, was identified indicating phosphate was bindingand sequestering available Zn and Zn pore water concentrations were decreased at levels of 90% or more.Laboratory experiments indicate organic amendments exhibit a limited effect and may hinder sequestra-tion of pore water Zn when mixed with apatite. The acute toxicity of the sediment Zn was evaluated withHyalella azteca, and bioaccumulation of Zn with Lumbriculus variegates. The survivability of H. aztecaincreased as a function of phosphate (apatite) loading rate. In contaminated sediment without apatite, nospecimens of H. azteca survived. The bioaccumulation of Zn in L. variegates also followed a trend ofdecreased bioaccumulation with increased phosphate loading in the contaminated sediment. The researchsupports an association between Zn speciation and bioavailability.

Keywords: zinc, remediation, phosphate, in situ immobilization, X-ray absorption spectroscopy

INTRODUCTION

As a result of anthropogenic inputs, many sediments in

coastal waterways contain metal concentrations that exceed

established threshold values for toxicity to biota and benthic

organisms. Consequently, each year many millions of cubic

yards of sediments are capped or dredged for environmental

cleanup in commercial and recreational waters (Crannel

et al., 2001). Due to the high cost of these approaches,

alternative in situ remediation strategies capable of reducing

biotoxicity in sediments have been considered to manage

risk (Porter et al., 2004). Of the methods implemented,

many utilize amendments that stabilize or sequester metals,

however, an understanding of what the in situ stabilization

mechanism is and how it affects bioavailability is not well

known (Scheckel et al., 2009). Further, the amendments

must ensure no further or potential long-term harm to

delicate sediment ecosystems.

To date, in situ remediation of metals with amendments

has largely been limited to soils. For example, soils

contaminated with lead (Pb), zinc (Zn) and cadmium (Cd)

have been treated in situ with phosphate (P) or biosolids

amendments (Basta et al., 2001; Basta and McGowan,

2004; Boisonn et al., 1999; Brown et al., 2004; Cao et al.,

2003; Laperche et al., 1996; O’Day et al., 1998; Ryan et al.,

2004; Ryan et al., 2001; Scheckel and Ryan, 2004;

Scheckel et al., 2005; Wang et al., 2001). The anticipated

decrease in metal bioavailability in P-amended soils is

attributed to both the rapid kinetics of metal sequestration

and the low solubility of metals complexed with P or

isomorphically substituted into the apatite structure. The

mechanism of metal immobilization by apatite in the

literature is varied and dependent on the metal. In the

case of Pb, sorption exhibits a low dependence on pH

and Pb sequestration appears to be limited by apatite

dissolution (Chen et al., 1997; Ma et al., 1994). For Zn,

however, the uptake of aqueous Zn is strongly related to

solution pH suggesting the dominant mechanism of removal

is sorption at the apatite surface, which may be a benefit in

saturated sediments compared to dryer soil environments

(Chen et al., 1997, Xu et al., 1994). In situ remediation of

sediments, however, has not been widely applied (Scheckel

Chemical Speciation and Bioavailability (2011), 23(3) 163

www.chemspecbio.co.uk

doi: 10.3184/095422911X13103699236851

et al., 2011). Particularly considering the harmful offset of

excess phosphate in a freshwater system, it is vital that

appropriate amendment products be utilized to limit

ecosystem impact. In addition, sediments offer a unique

set of environmental variables; they are subject to ambient

conditions and a lack of process control after the application

of amendments (Renholds, 1998). Sediments are also

subject to either complete saturation or intermittent fluctua-

tions in saturation and oxicyanoxic conditions. As a general

rule, in situ methods also face the challenge of exhibiting

lower efficiency levels than ex situ methods, thus limiting

their wide acceptance (Renholds, 1998).

The site of this sediment study is the Indian Head Naval

Surface Warfare Center (IHNSWC) located 25 miles south

of Washington, D.C. in Charles County, Maryland, USA.

Historical information indicates a Zn recovery furnace was

built in 1928 near the shore of Mattawoman Creek at the

northeast border of the IHNSWC (Hill, 2004). Mattawoman

Creek is a tributary to the lower Potomac River and is

located in a tidal freshwater-estuary ecosystem. This facility

was used to recover Zn under the Navy metal-recycling

program during World War II and a review of station maps

indicate the recovery building was dismantled in the 1950s.

Presently, the site is characterized by a lack of vegetation

and active surface erosion. In 1993, a soil survey conducted

near the site of the Zn recovery furnace identified Zn at

7350 ppm in the near shore sediment (Hill, 2004). In 2000,

additional soil and sediment samples were collected and

identified labile Zn in the sediment pore water exceeding

25 mg L� 1 (SAIC, 2001). As a result of the high concen-

trations, Zn was identified in a Toxicity Identification

Evaluation report as the metal of concern (SAIC, 2001).

Based on site findings regarding the levels of Zn in the

sediment and a desire to avoid dredging, this study

evaluated the in situ stabilization of Zn with natural

amendments to reduce biotoxicity and bioaccumulation to

acceptable levels in the benthic community. The objectives

of this study were to: (i) evaluate Zn speciation in natural

sediments prior to and after addition of amendments, (ii)

perform biological assays to determine the toxicity and

bioaccumulation of the sediment Zn prior to and after

addition of amendments, and (iii) link Zn speciation or

the change in Zn speciation to benthic organism bioavail-

ability based on survivability and bioaccumulation studies.

METHODOLOGY

Field plots

The experimental design was based on test plots measuring

3 m by 3 m by 0.2 m located in an intertidal zone at the

shore of Mattawoman Creek. Two control plots located

upstream of the contaminated site were used to determine

the effect of the apatite amendment on benthic organisms

and phosphate levels in the absence of Zn. Active plots,

located adjacent to the vegetative dead zone down-slope of

the former Zn smelter, were used to monitor the effect of

apatite on Zn sequestration and speciation in the sediment.

Apatite loading rates were chosen on the hypothesis of

stoichiometric Zn phosphate formation. Stoichiometry indi-

cates the ideal ratio of P to Zn, based on the formula for of

hopeite [Zn3(PO4)2(OH, F, or Cl)]. This is the same ratio

required for, and shown to work for Pb in soils (Cao et al.,

2002; Ma et al., 1994). In addition to apatite, all amended

plots received a 15% by mass addition of Orgro1 biosolids

material. Orgro1 is a soil conditioneryfertilizer and does

not exceed the USEPA Alternative Pollutant Limits for

metals. The Orgro1 biosolids will act as both a possible

metal sequester in the sediment and a natural enhancer to

support vegetative growth (Farfel et al., 2005).

Excluding pore water, the density of the sediments was

estimated at 2.65 g cm� 3. The volume of sediment in each

plot of (36360.2) m3¼ 1.8 m3 or 1.86106 cm3 or

4.776106 g. The phosphate rock used in this study is

20% phosphate as P2O5, or 8.7% phosphorous by mass.

Based on the mass of P in apatite and the effective or

accessible mass in the test plots, a 1% P loading will require

318 kg of apatite and a 0.5% P loading will require 159 kg.

Prior to the application of apatite and Orgro1, plots were

staked out, separated with silt curtains and mechanically

mixed with a garden rototiller to ensure the homogeneity of

the site and sampled. All sampling was performed by

collecting a composite of twelve subsamples from each

plot and homogenizing the sample by mixing. After

sampling, apatite and Orgro1 were spread over the

surface of the plots according to the desired loading rate

at low tide and mechanically mixed into the sediment with

the rototiller. The rototiller was used to mix the amend-

ments to a depth of approximately 20 cm, with mixing

accomplished by running the machine in three directions

across the full width of each plot.

Bioassays

Composite sediment samples from each of the five plots at

the shore of Mattawoman Creek were collected 4 months

after amendment installation during low tide to a depth of

approximately 5 cm and prepared for bioassay tests

according to the procedures and methods outlined by

USEPA guidance documents (USEPA, 2000). The proce-

dure measures chronic toxicity through direct sediment

exposure to the freshwater amphipod, Hyalella azteca,

over a 28-day period and bioaccumulation of Zn in the

tissue of Lumbriculus variegates, also during a 28-day

period. For the acute toxicity test with H. azteca, a

composite sediment sample from each plot of sufficient

volume (approximately 26.5 litres) was collected, mixed

and split between 12 tanks with 10 organisms per tank.

Therefore, 12 mean survival responses and 12 mean growth

responses were obtained to evaluate each sediment sample

for each of the five plots. For the bioaccumulation test using

L. variegates, a preliminary four-day toxicity screening test

was performed to determine if L. variegates would survive

and exhibit normal behaviour, e.g. burrow into the sediment

(USEPA, 2000). Based on the survival and normal beha-

viour of the L. variegates worms in the preliminary test, a

28-day bioaccumulation study was performed as detailed in

USEPA guidance documents (USEPA, 2000). Briefly,

164 Speciation and bioavailability of zinc

approximately 5 g (wet-weight) of L. variegates was

measured and recorded before being placed into replicates

glass jars, 12 replicates for each of the five sediment test

plots. During the test period, renewal of the overlying water

was accomplished twice daily by automatic flow controls.

At the end of the 28-day period, L. variegates was allowed

to depurate for 24 hours in clean water before being

analyzed for Zn tissue concentrations (USEPA, 2000).

Ammonia (SM-4500-NH3), temperature (SM-4500), pH

(SM-4500-Hþ ), dissolved oxygen (EPA Method 360.1),

hardness (EPA Method 130.2), alkalinity (EPA Method

310.1) and conductivity (EPA Method 120.1) were

measured during the test periods for both the survivability

and bioaccumulation tests to account for these potential

effects on the organisms in addition to the toxicity of Zn.

Analysis of variance, or Kruskal Wallis for non-normal

response, was used to test for differences between bioassay

results for the treatments. A control sediment consisting of

quartz sand was used to observe health affects in the

absence of sediment for both the H. azteca and L. varie-

gates bioassay tests.

Bench-scale study of organic amendments

To evaluate the influence of organic amendments on zinc

sequestration by apatite and study potential benefits of

higher phosphate loading rates, two additional test plots

were amended with 3% and 5% P as apatite and allowed to

equilibrate for 12 months without addition of any organic

amendment. At 12 months, sediment was collected from the

top 5 cm of each plot to be consistent with previous

sampling efforts. Collected sediments were sent to the

Battelle Marine Sciences Laboratory in Sequim, WA to

perform the organic amendment and laboratory tests. Two

commercially available organic amendments were chosen to

add to the collected apatite-amended sediments to evaluate

the effect of organic amendment addition. The two amend-

ments were a biosolids sold commercially under the name

ComPro2, but equivalent to the Orgro used in the field

amendments, and an organic leaf mould, sold commercially

under the name LeafGro2.

Three different organic treatments (no organic amend-

ment, 15% biosolids (ComPro2), 15% leaf mould

(LeafGro2)) were applied to sediments from each plot

(0%, 3%, and 5% phosphate), yielding a total of nine

different treatments. Zinc and ortho-phosphate (ortho-P) in

sediment pore water were measured via inductively coupled

plasma mass spectrometry (ICP-MS; Perkin Elmer) in

homogenized samples from the three test plots at test

initiation and in all treatments at Weeks 4 and 10. A total

of six replicates were set up for the ‘‘no organic amend-

ment’’ treatments, with two replicates taken down at each

sampling interval for pore water analysis. A total of four

replicates were set up for the organic amended sediments,

with two replicates taken down at Week 4 and two

replicates taken down at Week 10 for pore water analysis.

Flow-through bench top study conditions were used to

mimic the laboratory bioassay with the amphipod

Leptocheirus plumulosus, except that no organisms were

added to the test containers (USEPA, 2000). Key conditions

of the test were water temperature maintained at 25�C,

salinity 20%, gentle aeration of the test containers, and

periodic exchange of the overlying water (50% of the

volume, 36yweek). The photoperiod was maintained at

16 h light, 8 h dark.

Sediment and zinc analysis

Composite sediment samples from each test plot before and

after amendment addition were collected by placing the

sediment in glass bottles without headspace, stored under

ice, and shipped to the USEPA National Risk Management

Research Laboratory in Cincinnati, OH. In the lab, the

sediment bottles were placed in an anaerobic chamber. A

portion of the sediment from each plot was freeze-dried and

lightly crushed with an agate mortar and pestle for spectro-

scopic analysis under anaerobic conditions.

Environmentally available metals in the sediment were

determined with EPA Method 3051 in triplicate. A

National Institute of Standards and Technology reference,

standard reference material1 2711, was used for quality

control. Concentrations of 26 elements from the sediment

digestions were determined using inductively coupled

plasma atomic emission spectrometry (ICP-AES). Matrix

matched (10%, Trace Metal Grade HNO3, Fisher Scientific,

Fairlawn, NJ) ICP-AES standards were prepared from

certified stock solutions.

Equilibration of sediment with 10 mM CaNO3 as a

background electrolyte in DI water was performed under

oxic and anoxic condition to observe the effect redox

conditions have on Zn desorption. A 1 : 100 solid solution

ratio was used. The pH was maintained at 7.0 with

automatic addition of 10 mM NaOH with a pH-STAT.

Anoxic conditions were attained by first degassing the

10 mM CaNO3 solution with high-purity N2(g) for several

hours before addition of the sediment. The N2(g) was

maintained during the experiment until oxic conditions

were selected for by ceasing N2(g) flow.

X-ray diffraction was carried out on freeze-dried sedi-

ment. Scans were completed between 5� and 80� with a step

of 0.02� s� 1 (X’Pert, Panalytical, The Netherlands).

Mossbauer spectra were collected in transmission mode

with a constant acceleration drive system and a 57Co

source. Samples were mounted in a top-loading Janis

exchange-gas cryostat and data was calibrated against an

a-Fe metal foil collected at room temperature. Spectral

fitting was done with the Recoil software package using

Voight based spectral lines with a fixed line width

(HWHM¼ 0.097 mmys) (University of Ottawa, Ottawa,

Canada).

Zinc K-edge (9659 eV) X-ray absorption and m-X-ray

fluorescence (m-XRF) spectroscopies were conducted at

beamline XORyPNC (Pacific Northwest Consortium

Collaborative Access Team) and MR-CAT (Materials

Research Collaborative Access Team ) at the Advanced

Photon Source at Argonne National Laboratory, Argonne,

IL. The electron storage ring operated in top-up mode at

7 GeV. Spectra were collected in both transmission and

Aaron G.B. Williams, Kirk G. Scheckel, Gregory McDermott, David Gratson, Dean Neptune and James A. Ryan 165

fluorescence mode with a 13-element solid-state detector at

room temperature. For each sample, a total of three to five

scans were collected and averaged. Data were analyzed with

the IFEFFIT software program (Ravel and Newville, 2005).

The results for the samples were compared with those from

synthesized minerals and mineral specimens acquired from

the Smithsonian Institute. All minerals were verified with

XRD before use as reference materials for assessment of Zn

solid-state speciation.

RESULTS AND DISCUSSION

Sediment from the shore of Mattawoman Creek is largely

composed of quartz sand, with small amounts of mica and

muscovite based on XRD pattern matching. Minor peaks in

the XRD pattern appear to be consistent with other

phyllosilicates such as kaolinite. The phases observed

with XRD were consistent through the size fractions and

did not identify any Zn bearing minerals. Zinc bearing

minerals were not observed due to low crystallinity or the

mass percentage is too low to be observed with this method.

Although no Zn bearing phases were identified with XRD,

chemical digestion (EPA Method 3051) of freeze-dried

sediment released high levels of Zn, levels as high as

47.6 g kg� 1 of dry sediment, or 4.8% by mass. The

distribution of Zn based on particle size shows the Zn is

concentrated in a bimodal pattern, with the highest concen-

trations observed for the silt and clay size particles

(538 mm) and for the larger sand grains (40.85 mm).

The larger sand grains exhibit an orange coating typical

of iron oxidation, which may aid in the adsorption or

precipitation of metals at their surface accounting for their

high Zn content. The relationship between Zn and the other

elements observed in the sediment based on the chemical

digest are shown in Figure 1. A clear relationship is

observed between available Zn, Fe and Al and to a

limited degree with S, with the strongest correlation

observed between Fe and Zn (Figure 1 inset). The correla-

tion of Fe and Zn is also observed spectroscopically with

m-XRF (Figure 2). A comparison of 2-dimensional m-XRF

plots for Fe and Zn in sediment thin sections show the Zn is

concentrated in areas of high Fe content (Figure 2). The

speciation of the Zn associated with the Fe in the m-XRF

images is consistent with Zn(OH)2 and Zn – Al layered

double hydroxides (LDH), with minor contributions from

Zn associated with ferrihydrite based on linear combination

fits of the data. A few Zn hot spots in the m-XRF images,

however, are not correlated with Fe (point A, Figure 2).

These spots are identified as smithsonite (ZnCO3) with

m-XAFS (Figure 2). The identification of ZnCO3 in the

sediment is expected. The residual slag material

surrounding the site of the former recovery furnace is

composed of zincite (ZnO) and ZnCO3 based on XRD

analysis. Because of Zn phytotoxicity, vegetation on the

hillside between the site of the Zn smelter and the creek is

barren and signs of active erosion are evident. Erosional

transport of the slag material, including ZnCO3, to the shore

of Mattawoman Creek is expected and likely accounts for

its presence in the XRF images; however, in situ formation

cannot be excluded.

To gain further insight into the behaviour of sediment Zn

and possible mineral phases present in the sediment, the

stability of the sediment was studied under variable redox

conditions. Under oxic conditions, or when contaminated

sediment was allowed to equilibrate with 10 mM CaNO3

open to the atmosphere, the aqueous Zn concentration

immediately increased in the first couple of minutes to 5 – 8

mg L� 1. After the initial desorption of Zn, the aqueous Zn

concentration slowly continued to increase over a period of

days eventually stabilizing around 20 mg L� 1 (Figure 3).

Similar to this response, contaminated sediment introduced

to deoxygenated 10 mM CaNO3 also underwent a rapid

release of Zn to the aqueous phase attaining levels of 5 – 7

mg L� 1. The aqueous Zn concentration, however, in

contrast to the oxic system maintained a constant value

5 – 7 mg L� 1 for several days. After the introduction of

oxygen to the anoxic reactor however, Zn is slowly released

over a period of days similar to the response observed in the

oxic system (Figure 3). The initial response of the sediment

in both oxic and anoxic water indicates a soluble or

adsorbed phase attains rapid equilibrium regardless of the

redox condition, with both systems attaining an aqueous Zn

concentration of 5 – 9 mg L� 1. A slow release of Zn after

the initial rapid release in oxic conditions however, suggests

Zn bound with reduced sulfur phases, such as Zn sulfide

(sphalerite) may also account for some Zn in the sediment.

The slow release of Zn under oxidizing conditions is

consistent with kinetics of redox limited reactions, such as

ZnS oxidation, compared to rapid changes observed for

adsorptionydesorption limited reactions that likely occur

when the sediment is initially exposed to a clean aqueous

environment. The rapid desorption of Zn from the sediment

has direct implications for Zn release during re-suspension

events that may occur during periods of high flow or from

dredging operations.


Figure 1 Elemental composition of Mattawoman Creek sediment

prior to amendment addition as determined with EPA method

3051. Inset: Mass relationship of Fe and Zn.

Preamendment sediment analysis

The distribution of Zn within the sediment based on particle

size is not uniform as shown in the sediment digest data in

Figure 1, however, XANES analysis of the sediment before

amendment addition indicates that regardless of the particle

size the speciation of Zn within the bulk sediment is largely

homogeneous (Figure 4). All sediment fractions contain

three identifiable Zn bearing phases: Zn(OH)2 [representing

zinc hydroxide and related zinc layered double hydroxide

(LDH) species], ZnS [sphalerite], and Zn – Fe oxide [repre-

senting zinc sorbed to the surface of an iron oxide]. The

bulk XANES spectra identify Zn(OH)2-like compounds as

the primary phase in the sediment accounting for 65 – 80%

of the total Zn. The other two phases, ZnS and a Zn – Fe

oxides phase, accounted for 3 – 20% and 8 – 20% respec-

tively. In addition, consistent with the chemical digest

results, the mid-size sediment fractions (125 – 250 and

250 – 425 mm) that exhibit low levels of available Fe also

exhibit lower Fe – Zn oxides concentrations based in linear

fits of the chi-space data.


Figure 2 m-XRF of Mattawoman Creek sediment thin-section identifying the Fe and Zn concentration profile and k-space spectra derived

from m-XAFS of point A and B.

Figure 3 Effect of oxicyanoxic conditions on desorption of Zn

from Mattawoman Creek sediment. Anoxic condition attained by

purging reactor solution with N2 gas.

The bulk spectroscopic analysis of the shoreline sedi-

ment did not identify the trace amounts of ZnCO3

identified with m-XAFS or the ZnO identified in the slag

material. The ZnCO3 was not observed in the bulk analysis

because of its sparse occurrence based on several 2-

dimensional scans of sediment thin sections. The other

Zn phase observed in the slag material, ZnO, is not

observed in either bulk or micro synchrotron analysis.

The ZnO may undergo complete dissolution before or

after transport to the shore of Mattawoman Creek, but its

absence in the creek sediment based on spectroscopic data

indicates it does not persist in the aqueous sediment

environment, for example, hydrolysis may convert the

ZnO to Zn(OH)2 and Zn – Al LDH phases.

The occurrence of Zn(OH)2 and Zn – Al LDH in the

sediment at the shore Mattawoman Creek, but not in the

Zn source, implies that it forms as a secondary phase as a

result of chemical precipitation or conversion. Based on the

solubility product (Ksp) of 3610� 17 for Zn(OH)2 and an

observed aqueous concentration on the order of 15 – 20 mg

L� 1, Zn(OH)2 solubility will be exceeded at pH values

around 5.6 to 5.8 without activity corrections. The forma-

tion of Zn(OH)2 phases, however, is likely enhanced due to

surface adsorption on sediment media, specifically with

regard to the Fe oxides, Fe oxide coated sands, and clay

minerals. The Fe phases observed in the sediment largely

represent poorly ordered Fe oxides with trace amounts of

hematite and Fe2þ containing clays from Mossbauer

spectroscopy (Figure 5) and XRD.

Post amendment sediment analysis

Post amendment analysis of Zn contaminated sediment

amended with apatite identified a change in Zn speciation

consistent with the formation of a zinc phosphate (ZnP)

species (Figure 6). The control plot in the contaminated

zone, which received no amendments, however, continued

to exhibit a XANES spectrum identical to the preamend-

ment plots indicating a positive response to amendment

addition (Figure 6). Linear combination fits of homogenized

samples from the apatite treated plots indicate the 0.5%

phosphate amendment plot consisted of 23 – 28% ZnP, 18 –

22% ZnS, and 51 – 56% Zn(OH)2-like phases, and the 1.0%

phosphate amendment plot consisted of 24 – 28% ZnP, 22 –

25% ZnS, and 48 – 52% Zn(OH)2-like phases (Figure 7).

Although the percentage of phosphate as apatite doubled

from the 0.5% to the 1.0% treatment plot, a significant

increase in ZnP was not observed, in fact, both plots exhibit

a maximum conversion of 28% Zn to ZnP based on the

spectroscopic analysis of the collected sediment. The

similarity in Zn speciation for the two plots may indicate

the available pool of Zn for phosphate complexation has

been exhausted, possibly due to limited aqueous Zn concen-

trations (486 and 123 mg L� 1 for 0.5% and 1.0% phosphate

treatments, respectively (Table 1)) or passivation of

Zn(OH)2 or apatite particle surfaces with ZnP coatings.

Otherwise, an increase in the ZnP species should be higher

in the 1.0% plots due to a doubling in the available surface

area and P mass. Apatite particles within the sediment were

visibly identifiable after the four months of treatment and

could be physically separated from other sediment particles.

XANES analysis of the crushed apatite particles identified

ZnP as the primary species associated with the apatite

particles indicating sorption of the Zn to apatite had

occurred (Figure 8).


Figure 4 XANES spectra of Mattawoman Creek sediment sepa-

rated by particle size.

Figure 5 Mossbauer spectra of Mattawoman Creek sediment.

Data collected at room temperature.

Wet chemical analysis of the pre- and post-amended

sediment plots also support a change in sediment Zn.

Pore water values indicated a significant decline in the

free or aqueous Zn after amendment addition (Table 1).

The plot with 0.5% P as apatite experienced a decrease of

5114 mg L� 1 Zn or 91% and the 1.0% P as apatite plot had

a 7177 mg L� 1 decrease, or 98% drop in pore water Zn

(Table 1). The control plot however, which did not demon-

strate a change in Zn speciation based on XANES analysis,

also experience a decrease of 3145 mg L� 1 Zn or 44%. The

unexpected decrease in Zn pore water in the control plot

indicates other influences, such as natural porewater fluxes

of Zn or artefacts from the development of the test plots (i.e.

disruption of sediment aggregation or preferential pore

water paths via rototilling), had an effect on the in situ

Zn concentrations. The aggressive mixing of the sediment

before preamendment sampling was intended to homoge-

nize the upper 15 – 20 cm of the sediment for the study, but

it may have altered the sediment Zn profile influencing the

mobility and speciation of the Zn through agitation or

disrupting redox boundaries and sediment particle distribu-

tions. It also demonstrates the difficulties and uncertainties

of studying in situ processes in natural systems as compared

to closed system laboratory investigations. To compound

the issues with pre- and post-sediment Zn concentrations,

we also observed a significant decrease in total Zn concen-

tration in the sediment in all three contaminated test plots

after the four months of treatment. All of the contaminated

plots experienced a decrease in total Zn mass (Table 1). A

decrease in total Zn on a per mass basis was anticipated due

to the addition of apatite and Orgro1 material, which

lowers the mass percent of original sediment in each plot,

however, this impact was estimated to be approximately

30% in the top 20 cm of the sediment. The observed

decreases in total Zn, however, are on the order of 40 –

80% again indicating other factors are influencing the Zn

distribution based on collected data. The low recovery may

also indicate a higher level of sediment heterogeneity than

anticipated from the composite samples. Historical

discharge data for Mattawoman Creek does not indicate

any unusually high discharge rates that may have trans-

ported or scoured the test plots (USGS, 2006). The altering

of the test plot density due to Orgro1 addition, however,

may have increased sediment porosity and enhanced Zn

transport in addition to increasing the possibility of

increased scour and transport of the less dense sediment.

The ortho-P pore water concentration was measured to

determine its potential affect on water chemistry. The

control plot in the uncontaminated zone with 0.5% P as

apatite increased from 0.02 mg L� 1 to 0.06 mg L� 1. The

0.5% P as apatite plot in the contaminated zone, however,

had a decrease in ortho-P due to the high level observed in

the pre-amendment sample, but the post-amendment sample


Figure 6 XANES spectra of Mattawoman Creek sediment from

apatite treated test plots compared to control plot without apatite

(no P). ZnS, ZnyFe(OH)3, and Zn(OH)2 standards shown for

comparison.

Figure 7 Linear combination fit of 1.0% apatite treated plot

XANES spectra.

had an ortho-P concentration in the range of 0.07 – 0.12 mg

L� 1 slightly higher than the equivalent control. The 1.0% P

as apatite test plot had the greatest increase in ortho-P

concentration with an increase from 0.05 mg L� 1 prea-

mendment to 0.15 – 0.22 mg L� 1 post amendment.

Bioassay

The mean survival rate of the benthic arthropod H. azteca

was determined in two uncontaminated, three contami-

nated, and one quartz-sand control sediment samples.

Each of the six conditions was completed in 12 replicate

reactors with 10 organisms per container as described in

the methods section. All geochemical variables measured

(temperature, pH, dissolved oxygen, hardness, alkalinity

and conductivity) were nearly identical in all trials during

the 28-day test period (Table 2). The only exception was a

significantly lower dissolved oxygen concentration in the

control sediment; however, the lower dissolved oxygen

concentration did not result in any observed decrease in

survivability. The effect of the apatite loading rates on

survivability and amphipod weight is presented in Table 3.

In the uncontaminated sediment, the mean survival rate

was greater than 90%. This value was statistically not

different from the mean survival rate observed in the

laboratory sediment controls consisting of clean quartz

sand, which was 88%. In the contaminated creek sediment,

which received no apatite or Orgro1, no specimens of

H. azteca survived the 28-day test period indicating the Zn

concentration was lethal to the amphipod. In contrast to the

lethality of the unamended sediment, the survival of

H. azteca in apatite-amended sediment was marked. The

contaminated sediment amended with 0.5% P as apatite

has a survival rate of 0 – 60% with a mean of 31%. A

doubling of the apatite loading rate to 1.0% P resulted in

an approximate doubling of the mean survival rate to 58%

with a range from 10 to 90%. The large variability in

survival rates among the application rates suggest a high

degree of variability of Zn in the sediment, however, an

increase in the survival rate did improve with increased

apatite loading rates. In addition to the survival rate, the

sub-lethal endpoint of growth was also evaluated. In all

studies that exhibited survival of H. azteca some growth of

varying degrees was observed and comparable to the

apatite loading rate. (Table 3).

The survival of H. azteca, as with other invertebrates is

observed to be a function of the free or aqueous Zn

concentration. Recently, a study demonstrated that

washing sediment of free or labile Zn increased the

survivability of Daphnia magna from less than 50% to

95% (Gillis et al., 2006). The concentration of Zn lethal to

H. azteca is unknown from our work, but the aqueous

concentration level toxic to D. magna has been studied. A

recent study demonstrated aqueous Zn was tolerated up to

a concentration of 170 mg L� 1 with a survival rate of 93%,

however, as the Zn concentration increased to 250 mg L� 1

and 340 mg L� 1 the survivability drastically decreased to

40% and 7%, respectively, after one week (Muyssen et al.,

2006). Considering the aqueous Zn concentrations

observed in the Mattawoman Creek sediment are typically

an order of a magnitude higher, the sediment is likely toxic

to many benthic organisms in addition to H. azteca and

D. magna.


Table 1 Zn pore water concentration and Zn sediment concentration observed before and after apatite addition to shoreline test plots

Pore water Zn (mg L� 1) Dry weight sediment Zn (mg kg� 1)

Condition Pre-amendment Post amendment Pre-amendment Post amendment

Control Not amended 8.63 8.9 62.4 390.5% apatite 1.95 1.96 75.6 140

Contaminated Not amended 7145 4000 35834 182060.5% apatite 5600 486 37239 217801.0% apatite 7300 123 47616 16368

Figure 8 XANES spectra of apatite particles collected from 1.0%

apatite test plot after four months exposure to contaminated

sediment. Apatite particles were separated from the sediment,

washed and crushed to a powder for analysis. ZnP and Zn(OH)2

standards are shown for reference.

The bioaccumulation of Zn in the tissue of L. variegates,

like the survivability of H. azteca, responded according to

the mass of phosphate applied to the sediment. All other

geochemical parameters other than the phosphate loading

rate were similar (Table 4). In unamended contaminated

sediment, the mean Zn tissue concentrations was 3070 mg

g� 1, with 0.5% P as apatite the Zn tissue concentration

decreased to 1610 mg g� 1, and at 1.0% P as apatite the Zn

tissue concentration further decreased to 654 mg g� 1. The

control plots in uncontaminated sediment had an accumula-

tion of 398 mg g� 1 in the unamended sediment and 384 mg

g� 1 in the 0.5% amended sediment. Again, the bioassay

demonstrates a relationship to the phosphate loading rate,

which was also shown to alter the Zn species in situ.


Table 2 Summary of water quality measurements for the 28-d Hyalella azteca chronic toxicity test

DissolvedTemperature pH oxygen Hardness Alkalinity Conductivity

Treatment (�C) (mg L� 1) (mg L� 1 CaCO3) (mg L� 1 CaCO3) (mS)

Min Max Min Max Min Day 0 Day 27 Day 0 Day 27 Day 0 Day 27

Target range: 22 24a No guidelinec 42.5 No guidelinec No guidelinec No guidelinec

20 26b

Control not amended 21 24 8.2 8.6 6.6 153 170 130 150 0.34 0.41Control 0.5% apatite 21 23 8.2 8.6 7.1 153 170 130 160 0.35 0.50Contaminated 0.5% apatite 21 23 8.1 8.4 7.0 153 170 125 140 0.31 0.47Contaminated not amended 21 24 7.9 8.5 6.7 136 170 125 140 0.34 0.44Contaminated 1.0% apatite 21 23 7.9 8.4 6.4 136 153 140 160 0.34 0.51Control Sediment 21 24 7.8 8.5 3.5 153 170 130 140 0.36 0.50

aDaily average for duration of test.bAllowable day-to-day variation.cMust not vary more than 50% during test.

Table 3 Mean survival and mean weight of Hyalella azteca exposed to Mattawoman Creek sediment as a function of contamination andamendment addition

Condition Survival mean(%)

Survival CV(%)

Weightyindividualmean (mg)

WeightyindividualCV (%)

Control Quartz sand 88 0.310 25Not amended 96 5 0.603 90.5% apatite 92 11 0.409 11

Contaminated Not amended 0 0 0 00.5% apatite 31 67 0.083 571.0% apatite 58 24 0.148 53

Table 4 Summary of water quality measurements for the 28-d Lumbriculus variegates bioaccumulation assay

DissolvedTemperature pH oxygen Hardness Alkalinity Conductivity

Treatment (�C) (mg L� 1) (mg L� 1 CaCO3) (mg L� 1 CaCO3) (mS)

Min Max Min Max Min Day 0 Day 27 Day 0 Day 27 Day 0 Day 27

Target range: 22 24a No guidelinec 42.5 No guidelinec No guidelinec No guidelinec

20 26b

Control not amended 22 23 7.9 8.4 3.3 170 204 150 160 0.34 0.47Control 0.5% apatite 22 23 7.7 8.1 2.7 170 238 165 300 0.36 0.76Contaminated 0.5% apatite 22 24 7.8 8.2 2.1 170 121 160 300 0.34 0.74Contaminated not amended 22 24 7.6 8.1 2.5 170 121 155 350 0.36 0.74Contaminated 1.0% apatite 22 23 8.0 8.3 4.0 170 187 175 200 0.34 0.40Control Sediment 22 23 8.4 8.6 5.1 170 170 175 160 0.34 0.38

aDaily average for duration of test.bAllowable day-to-day variation.cMust not vary more than 50% during test.

Influence of bench-scale organic amendments

Twelve months after initiating the 3% and 5% phosphate test

plots in the absence of organic amendments sediment was

collected and evaluated for pore water zinc and ortho-P.

Zinc pore water concentrations decreased 95% and 94% for

the 3% and 5% test plots respectively, and the control plot

exhibited a 26% decrease. The decrease in the control plot

is consistent with the decreases observed for prior in situ

test plots in this study. Pore water ortho-P levels at the end

of one-year ranged from 0.025 mg L� 1 with no apatite, to

0.04 mg L� 1 at 3% phosphate and 0.055 mg L� 1 at 5%

phosphate (apatite) treatments indicating a release of

soluble phosphate. The one-year aged sediments were

then mixed with biosolids or leaf mould and aged for 10

weeks (Table 5).

At the end of the 10 week equilibration period, the

control reactors without apatite or organic amendment

exhibited a Zn pore water concentration of 1212 mg L� 1,

a reduction from the 8851 mg L� 1 observed just prior to 10

week study. An equivalent decrease was also observed

(1378 mg L� 1 with 15% biosolids, 1152 mg L� 1 with

15% leaf mould) in reactors amended with biosolids and

leaf mould but no apatite indicating the organic amend-

ments had minimal or no significant influence on Zn pore

water concentrations in the aged sediment. It is noted the Zn

pore water concentrations decreased in the control reactor in

the absence of any potential sequestering agents and may

have been influence by handling and mixing during reactor

preparations.

In reactors receiving sediment treated with 3% and 5%

phosphate as apatite, Zn pore water concentrations at one

year were much less than the control (8851 mg L� 1) at

743 mg L� 1 and 616 mg L� 1, respectively, even though all

test plots exhibited initial zinc pore water concentrations of

10,800 to 13,600 mg L� 1. The apatite only reactors

continued to exhibit a decreased in pore water zinc over

the 10 week period of the organic amendment study, with

both the 3% and 5% phosphate reactors at 433 mg L� 1 at 10

weeks. In the 3% and 5% phosphate as apatite reactors

receiving organic amendments a net increase in Zn pore

water was observed. Both the 3% and 5% reactors with

organic amendment exhibited approximately twice the Zn

pore water compared to the 3% and 5% reactors in the

absence of an organic amendment. The data appear to

indicate that the apatite amendments perform better in the

absence of organic amendments, or at the very least, that

organic amendments do not enhance the sequestration of

zinc in the period evaluated in this laboratory study.

Analyses were conducted to characterize the soluble zinc

composition of both the biosolids and leaf mould materials.

The two materials were similar in zinc composition, with

soluble concentrations ranging between 52 and 79 mg L� 1,

and not high enough to account for the differences observed

between the organically amended and non-organically

amended plots. Potential undesirable affects of the organic

amendments are release of trace metals that may complete

with Zn for soluble or reactive phosphate or loss of

phosphate directly to the organic amendments.

Pore water samples were also analyzed for ortho-P to

determine if the higher loading of apatite or organic

amendments resulted in increased phosphate production.

Phosphate production is of concern in aquatic environments

since excess phosphate leads to eutrophication of aquatic

systems. A summary of the ortho-P results is presented in

Table 6. In reactors amended with leaf mould, pore water

ortho-P levels increased 30% for leaf mould only, and by a

factor of 3 when mixed with apatite compared to the no

organic controls, which exhibited an average pore water

ortho-P level of 0.15 mg L� 1. The biosolids reactors

exhibited an increase of pore water ortho-P of approxi-

mately 7- to 10-fold higher compared to the no apatite

controls. In both organic amendment trials, pore water

ortho-P was higher in reactors with apatite.

At the conclusion of the 10 week bench-top study,

samples of each of the treatment sediments were also

examined by XANES to identify Zn speciation. Table 7

shows the XANES LCF results. The non-apatite, non-

organic amendment sample was determined to have about

27% zincite and 73% Zn(OH)2-like phases (Zn(OH)2 and

Zn – Al LDH) with no measureable zinc phosphate compo-


Table 5 Zn pore water concentrations as a function of apatite loading and organic amendment type

Apatite exposure timeline? 0 weeks 52 weeks 56 weeks 62 weeks

Organic amendmentexposure timeline? 0 Weeks 4 weeks 10 weeks

Treatment Apatite 0%, no organics 12,000 8,851 2,100 1,212Apatite 0%, with 15% biosolids 1,347 1,378Apatite 0%, with 15% leaf mould 1,080 1,152

Apatite 3%, no organics 13,600 743 492 433Apatite 3%, with 15% biosolids 881 880Apatite 3%, with 15% leaf mould 1,047 865

Apatite 5%, no organics 10,800 616 470 433Apatite 5%, with 15% biosolids 952 879Apatite 5%, with 15% leaf mould 947 727

Units: mg L� 1.

nents. The biosolids and leaf mould amendments without

apatite observed the presence of zinc phosphate (34 – 41%,

respectively) resulting in a lower proportion of zincite and

Zn(OH)2-like phases. When apatite was present without an

organic amendment, zinc phosphate accounted for 33% of

the total zinc phases during the 10 week trial with notice-

able reduction for zincite. The addition of biosolids and leaf

mould to the 5% P as apatite amendment resulted in a 10%

increase of zinc phosphate (43%) in comparison to apatite

alone, but in this case the Zn(OH)2-like phases were

significantly reduced in the high organic system.

CONCLUSION

The response of H. azteca and L. variegates to the

phosphateyorganic amended sediments and spectroscopic

evidence demonstrating a conversion of in situ Zn specia-

tion suggest Zn can be stabilized with a corresponding

decrease in bioavailability. Metal toxicity, specifically for

Zn, is associated with the free or aqueous pore water

concentration (Gillis et al., 2006; Lock and Janssen,

2003), which was reduced based on the pore water observa-

tions. The chemical transformation of Zn(OH)2 and Zn – Al

LDH, which may act as a buffer or source of aqueous Zn to

ZnP may decrease the availability Zn due to the low

solubility of ZnP. Overall, addition of apatite to the Zn

contaminated sediments resulted in a decrease of pore water

Zn concentrations, an increase in benthic invertebrate

survival and growth, and a decrease in benthic invertebrate

bioaccumulation corresponding to the rate of phosphate

applied and an observed change in Zn speciation.

Although none of the amended plots performed to the

level of uncontaminated plots after four months of treat-

ment, the results indicate a positive response and support in

situ treatment as a potential remediation strategy for metals

in sediments. The affect of longer equilibration periods,

different phosphate application rates, and apatite particle

sizeysurface area are currently being studied to enhance the

results of this study.

ACKNOWLEDGEMENTS

The U.S. Environmental Protection Agency through its

Office of Research and Development funded and managed

a portion of the research described here. It has not been

subject to Agency review and therefore does not necessarily

reflect the views of the Agency. No official endorsement

should be inferred. This work was completed by the

USEPA and Neptune and Company under a joint U.S.

Navy approved Quality Assurance Project Plan. Some

samples, data collection, and analysis were completed by

or directed by Neptune and Company. PNCyXOR facilities

at the Advanced Photon Source, and research at these

facilities, are supported by the US Department of

Energy – Basic Energy Sciences, a Major Facilities

Support grant from NSERC, the University of

Washington, Simon Fraser University and the Advanced

Photon Source. Use of the Advanced Photon Source is also


Table 7 Zn speciation by linear combination fitting as a function of apatite loading and organic amendment type at the end of a10-week laboratory bench-top study

Amendment Zinc phosphate(%)

Zincite (zinc oxide)(%)

Zn(OH)2

(%)ZnAlLDH

(%)

0% Apatite, no organics 0 26.9 19.1 54.00% Apatite, biosolids 34.4 20.6 0 45.00% Apatite, leaf mould 41.2 21.5 0 38.15% Apatite, no organics 33.3 18.1 0 48.65% Apatite, biosolids 42.6 37.8 0 19.65% Apatite, leaf mould 43.7 30.9 0 25.4

Table 6 Ortho-phosphate pore water concentrations as a function of apatite loading and organic amendment type

Apatite exposure timeline? 0 weeks 52 weeks 56 weeks 62 weeks

Organic amendmentexposure timeline? 0 Weeks 4 weeks 10 weeks

Treatment Apatite 0% , no organics 0.02 0.03 0.13 0.15Apatite 3%, no organics 0.05 0.04 0.09 0.16Apatite 5%, no organics 0.01 0.06 0.17 0.14Apatite 0%, with 15% biosolids 0.02 0.03 1.71 1.07Apatite 3%, with 15% biosolids 0.05 0.04 2.54 1.30Apatite 5%, with 15% biosolids 0.01 0.06 3.00 1.50Apatite 0%, with 15% leaf mould 0.02 0.03 0.40 0.21Apatite 3%, with 15% leaf mould 0.05 0.04 0.59 0.49Apatite 5%, with 15% leaf mould 0.01 0.06 0.80 0.45

Units: mg L� 1.

supported by the U.S. Department of Energy, Office of

Science, Office of Basic Energy Sciences, under Contract

DE-AC02-06CH11357. MRCAT operations are supported

by the Department of Energy and the MRCAT member

institutions. We wish to thank Mr. Andrew Gutberlet,

Environmental Engineer with the Naval Facilities

Engineering Command Washington 1314 Harwood Street,

SE Washington Navy Yard, DC 20374, for supplying the

bioassay and chemistry data results, and access to the Indian

Head site sediments.

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