Section C: HM

SECTION E

METAL BIOACCUMULATION & TISSUE BIOMONITORING

CHAPTER E1: METAL BIOACCUMULATION INTRODUCTION

The many sources of aquatic pollution in the highly urbanised area of south-eastern

Sydney include discharges from commercial, industrial and waste disposal sites,

sewerage overflow, leaching from former waste or industrial sites (many in unknown

locations), and general stormwater run-off (EPA 1996). Deliberate dumping and

accidental spillage may also be significant in the area due to the extensive industry

present (EPA 1996). Major contributors have been metal industries, chemical

manufacturers, users of chemicals (e.g. tanneries), and fuel producers (EPA 1996).

I compared metal accumulation (and associated reproductive effects, Section F)

between turtles from four sites in this highly industrialised area of south-eastern

Sydney, and turtles from four sites in national park areas located just to the south of

the city.

E1.1 Biomonitoring

To assess the risk of metal contamination on biota, inorganic components of the

environment (e.g. air, water, sediment) are often measured. Disadvantages of this are

that samples give only a snapshot view of conditions that may fluctuate over time, and

each of the environmental components represents only one pathway of possible

exposure for an animal. In contrast, biological monitoring encompasses exposure

from different environmental sources, different routes (inhalation, ingestion, and skin

contact), and over time (Gerhardsson & Skerfving 1996). Biological monitoring

provides a more meaningful estimate of metal contamination impacts than measuring

the inorganic environment, as long as there is a high correlation between metal

concentrations in the sampled tissue and the appearance of adverse effects (Camner et

al. 1979). Biological monitoring has been restricted by the limited performance of

analytical techniques, but inductively-coupled plasma mass spectrometry (ICPMS), as

used in the current study, now allows for analysis of low concentrations in small

sample volumes and determination of many elements simultaneously (Gerhardsson &

Skerfving 1996).

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Ideally the risk of adverse effects of a metal contaminant to an animal would be

assessed by measuring the concentration of the metal in the most sensitive organ, but

this would usually be fatal, and other tissues which are able to be sampled non-

lethally can be substituted and metal concentration in the critical organ subsequently

estimated (Camner et al. 1979). To use this method, the relationship between the

concentration of the metal in the sampled tissue, and the concentration of the metal in

the tissue of interest must be established, a correlation that has rarely been examined

for reptiles (Linder & Grillitsch 2000).

Many animal tissues and products have been tested for use in non-lethal

biomonitoring of metal exposure, including antlers (Kierdorf & Kierdorf 1999),

osteoderms (Twining et al. 1999), teeth (Rabinowitz et al. 1989), hair (Rivai 2001),

faeces (Fitzner et al. 1995, Eeva & Lehikoinen 1996), urine, blood (Drasch et al.

1997), eggs (Burger & Gibbons 1998), feathers (Janiga et al. 1990, Llacuna et al.

1995), scales (Kaur 1988), and nails (Nieboer & Fletcher 1996). Turtles are unique in

possessing an external bony shell, which means that metals in bone may be safely and

non-lethally sampled in these animals. It has been suggested that aquatic snakes are

useful bioindicators as they have relatively small home ranges, can be compared over

multiple habitats within a limited area, and can be maintained and used in a laboratory

settings (Campbell & Campbell 2001). These factors also apply to Australian

freshwater turtles, which have a much wider geographic distribution than Australian

aquatic snakes, and, unlike aquatic snakes, are found in Sydney (Cogger 2000).

Although home ranges of Australian turtles are not well known, E. macquarii can be

attached to a limited geographic location (Goode & Russell 1968) and, although

significant overland distances may be travelled by C. longicollis (Section A.1.3.4),

site fidelity may be high at some locations (Section D3.1.4, Goode & Russell 1968).

For animals to be useful bioindicators, they need to be able to survive in areas of high

pollution. Turtles show a high tolerance to pollution (Meyers-Schöne & Walton

1994), and populations can persist in environments severely polluted with metals

(Albers et al. 1986). Turtles may also be the largest or only vertebrate biomass in

some waters (Steven Emerton pers. comm.), especially at highly polluted sites (C.

longicollis, Section D2.2.5; Mauremys caspica, Yawetz et al. 1983). Additionally,

turtles are potentially useful biomonitors as they are long-lived, widespread and

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common, easy to capture, have a variety of diets, and live in a variety of habitats

(Meyers-Schöne & Walton 1994).

Lipophilic organic compounds such as DDT and PCBs are of particular concern for

top carnivores such as C. longicollis as they biomagnify through foodchains (Suedel

et al. 1994). Although metals transfer through aquatic food webs, biomagnification

only occurs for the limited number of metals that may be converted to organic forms

(e.g. Hg, As, Pb) (Suedel et al. 1994). Even with contaminants that do biomagnify

through aquatic food chains, there tends to be a 10-fold or less increase in

concentration from prey to predator, compared to the 100- or 1000-fold increases that

are reported for nonaquatic food webs (reviewed in Suedel et al. 1994). Also, reptiles

do not generally biomagnify metals to the same degree as other vertebrates at a

similar level of the food chain (Linder & Grillitsch 2000).

E1.1.1 Bone

The main sampling tissue for metal analysis during this study was the bone of the

marginal carapace. Turtle (and crocodile) bone tissues have a compact outer cortex,

surrounding a central medulla which is composed of cancellous trabeculae (Enlow

1969). Reptilian bone is cellular and usually vascularised (Enlow 1969), with many

small blood vessels within the bones of the carapace (Cann 1998), although the outer

cortex may be nonvascular (Enlow 1969). The chelonian shell is composed of a series

of bony plates, joined in a zigzag pattern of sutures, which support a germinative layer

responsible for producing the keratin which forms the overlying scutes (Alderton

1993). In Chelodina and Emydura species there are 11 pairs of marginal bones around

the outer edge of the carapace and these are overlaid with twelve pairs of marginal

scutes and one nuchal scute (Legler & Georges 1993), the joins of which do not

correspond to the joins of the bony plates (Alderton 1993). In cross-section, the

middle of the shell bones are spongy and contain many circular cavities of different

sizes, while the outer compact lamellar bone contains fewer vascular channels

(Zangerl 1969).

The mineral phase of bone is largely hydroxyapatite, which has the general formula

Ca10(PO4)6(OH)2. Thus, Ca is the major mineral of the bone matrix, but Na and K are

also present, along with essential trace metals such as Mg, Zn, Cu, and Mn that are

185

required for normal growth and remodelling of bone (Bhattacharyya et al. 1996).

Xenobiotic metals are also incorporated into bone, often for long periods (Tsuchiya

1979), where they may also cause bone abnormalities (Eeva & Lehikoinen 1996).

The histology of reptilian bone is very different to that of the better-studied

mammalian bone (Currey 1984), and turtles have much more complex and varied

bone histology than any of the other reptiles (De Ricqlès 1976). The carapace contains

an inner spongy layer limited by outer cortical layers of compact bone, and may have

little vascularisation, or contain numerous primary vascular canals.The histological

structure of the dense cortical sheets also varies greatly according to their anatomical

position, and from species to species (De Ricqlès 1976). Long bones show an

extensively developed inner cancellous endosteal region. Some cortical areas are

formed only of circumferential lamellae with little or no vascularisation, but as with

carapace bone, other areas of the same bone may contain numerous primary vascular

canals (De Ricqlès 1976).

Bone remobilisation is of interest as, along with Ca, potentially toxic metals bound in

the bone will be released during this process (Bhattacharyya et al. 1996). This may

lead to adverse affects on the individual or, for females, these metals may be

incorporated into eggs with potential ill-effects on the developing embryo. For

Australian freshwater turtles it is not known from which bone type Ca and other

metals are mobilised for egg formation. Limb bones have been indicated as the source

in some species (Sternotherus odoratus, Edgren 1960; Trachemys scripta, Suzuki

1963), but it is more likely that both limb and shell bone is involved (Magliola 1984).

Areas of bone remobilisation in reptiles are visible as Haversian canals (Twining et al.

1999). Resorption of bone from the walls of primary vascular canals leads to canal

enlargement, with the subsequent deposition of concentric lamellae within these

eroded spaces resulting in the formation of the secondary Haversian systems (Enlow

1969). Haversian canals are not found at all in lizards or snakes but are found in

localised areas of the bone cortex in turtles and crocodiles (Enlow 1969). They are

distinguishable from the primary canals which have not been remodelled by the

presence of the secondarily formed concentric lamellae, which are interspersed by

cement lines (Enlow 1969).

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As well as serving as a reservoir for toxic metals that may damage other tissues, the

bone itself is susceptible to metal attack. Both essential and xenobiotic metals have

specific and dose-dependent effects on bone cell function, including on the

relationship between the bone-forming osteoblasts and the bone-resorbing osteoclasts

(Bhattacharyya et al. 1996). The main osteotoxic metals are Cd, Pb, and Al, although

elevated levels of essential trace metals (e.g. Cu, Mg, Zn) may also be osteotoxic

(Bhattacharyya et al. 1996). Metals may also indirectly affect bone mineral

metabolism, e.g. by causing renal damage that leads to bone loss and increased bone

fragility (Friberg et al. 1979).

E1.1.2 Blood

Blood was also sampled from wild-caught turtles in this study. Whole blood rather

than plasma or serum should be analysed as metals stick to blood cells and proteins

(Dessauer 1970), and may be at highest concentrations within the erythrocytes

(Friberg et al. 1979). After exposure, animals concentrate metals in tissues that can

rapidly bind them, prior to being redistributed in the blood to longer term storage

tissues, so increases in blood metal concentration may be transitory. For example, the

majority of lead (90%) is stored in the bone where it has a halftime of 20 years,

whereas it has a halftime of only 20 days in the blood and soft tissues (Tsuchiya

1979). Thus, comparing changes in blood metals with changes in bone metals may

help elucidate whether exposure is current and chronic or historical.

Blood has been successfully used to monitor metal levels in sea turtles, but does not

always reflect high concentrations in other tissues (Kenyon et al. 2001).

E1.2 Aims

Current knowledge on the accumulation and effects of metals in reptiles is sparse, yet

metal pollution is a concern for all aspects of reptilian health and reproduction in

natural populations. Prior to this study, no work on metal accumulation or its effects

has been conducted on freshwater turtles in Australia. Worldwide, the only in-depth

studies have been conducted in North America on cryptodiran species (e.g. Bonin et

al. 1995, Bishop et al. 1998). Most other published results describe only general

clinical symptoms from opportunistic sampling of accidentally intoxicated turtles (e.g.

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Bury 1972, Borkowski 1997), or metal concentrations from a small number of tissues

(and often of turtles) with no examination of physiological affects (e.g. Beresford et

al. 1981, Davenport & Wrench 1990, Bonin et al. 1995, Burger & Gibbons 1998,

Gordon et al. 1998, Caurant et al. 1999, Anan et al. 2001).

My aims are to determine:

• which metals accumulate in Australian freshwater turtles (C. longicollis, E.

macquarii, and to a limited extent El. latisternum)

• which tissues the metals accumulate in, how concentrations in internal tissues

compare to those of carapacial bone, and whether internal organs have metal

concentrations of concern

• the concentration of metals in turtle tissues compared to the environment

• age (size)-related changes in metal accumulation

• adult differences in metal concentration related to sex, or egg production

• variations in metal concentrations between species

• the impact of urbanisation on metal accumulation

• Results will indicate whether non-lethal carapace or blood (or egg, Section F)

sampling and analysis is an effective biomonitoring technique for detecting toxic

metal exposures in these animals.

With the complete absence of any other studies tackling these issues in Australian

freshwater turtles, it is intended that this study will indicate more pressing areas for

future research.

Null Hypotheses

1. Metal concentrations in carapacial bone and blood (non-lethal sampling sites) are

independent of metal concentrations in internal organs. Rejection of this hypothesis

would provide evidence that carapacial bone and blood may be useful estimators of

metal concentrations in other tissues.

2. Metal concentrations in carapacial bone do not increase with turtle size. Rejection

of this hypothesis would provide evidence that bioaccumulation of metals occurs in

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bone. Non-rejection of this hypothesis means that metal concentrations in carapacial

bone may reflect recent environmental conditions rather than historical exposure.

3. Within sites, adult male and female turtles do not differ in their concentrations of

metal in carapace or blood. Rejection of this hypothesis suggests that the sexes are

exposed to different metal loads or that they have different metal storage capacities. If

so the sexes would have to be treated separately for other analyses.

4. Within sites, Chelodina longicollis and Emydura macquarii do not have different

concentrations of metal in carapace. Rejection of this hypothesis suggests that the

species are exposed to different metal loads or that they have different metal storage

capacities. If so the species would have to be treated separately for other analyses.

Non-rejection of this hypothesis permits pooling of data from both species for other

analyses.

5. Waterbodies in heavily urbanised environments do not have different

concentrations of metals in water or sediment to those in nearby national park areas.

Rejection of this hypothesis would allow testing of whether turtles inhabiting urban

waterbodies with high levels of metals have correspondingly high levels of metals in

carapace (Hypothesis 6a). Non-rejection of this hypothesis, but the presence of

significant differences among waterbodies leads to Hypothesis 6b.

6a. Turtles from heavily urbanised environments do not have different carapace metal

concentrations to those from sites in nearby national parks.

6b. Metal concentrations in turtle carapace are not correlated with metal

concentrations in the waterbodies from which they came.

Rejection of these hypotheses would provide evidence that turtles may have metal

concentrations that reflect the surrounding water quality, and hence that metal

concentrations could affect turtle health. Non-rejection of these hypotheses suggests

that turtles are not accumulating metals from their environment in any consistent

manner.

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CHAPTER E2: METAL BIOACCUMULATION METHODS

E2.1 Outline

Freshwater turtles (Chelodina longicollis, Emydura macquarii and Elseya

latisternum) were captured in lagoons at four urban (assumed to be polluted) and four

national park (assumed to have minimal pollution) sites. A sewage sludge lagoon was

opportunistically sampled as it represented a severely polluted environment. Water

and sediment samples were collected at each site. Turtles were weighed, measured, a

blood sample taken from the jugular vein, and a bone sample taken from the rear of

the carapace using a hacksaw. For reproductive studies (Section F), gravid females

were removed to the laboratory and egg-laying induced. Six eggs from each clutch

were frozen for later analysis, with the remaining eggs being incubated to hatching.

Hatchlings were weighed, measured and examined for external deformities.

E2.2 Sites

E2.2.1 General Location

Four urban lagoons in south-eastern Sydney, and three lagoons and one low-flow

creek located within national parks 30 km south of the city centre were sampled. The

urban lagoons were in the suburbs of Centennial Park (Urban 1), Eastlakes (Urban 2),

Botany (Urban 3) and Rockdale (Urban 4) (Figure E2.1; points U1-4), in densely

populated areas with a long history of industrialisation and subject to stormwater

runoff from roads. The non-urban waterbodies were Lake Toolooma (Park 1) in the

Heathcote National Park, and Kangaroo Creek (Park 2), Jibbons Lagoon (Park 3) and

Marley Lagoon (Park 4) in the Royal National Park (Figure E2.1; points P1-4).

Although not a lagoon, the water in Kangaroo Creek is usually still, only flowing

during heavy and prolonged rain (Royal National Park Service pers. comm.). All sites

are within 8 km of the coast and of similar altitude, latitude and longitude (Section

E2.2.2). Isolated lagoons were chosen as the lack of water transit would result in more

stable metal levels, and the movement of turtles in and out of sites would be

minimised. The sewage sludge lagoon incorporated into the Immune Study (Section

D2.2.5) was also sampled.

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The eastern suburbs coastal system incorporating the urban sites sampled here was

originally dominated by freshwater marsh habitat, and a characteristic array of heath

and shrubs known as Eastern Suburbs banksia scrub (Benson & Howell 1990, Keast

1995). The water supply of the larger ponds attracted industries such as tanneries and

wool-scouring works from the late 1800s, with swampland later reclaimed for housing

and airport development, and the Eastern Suburbs banksia scrub also diminishing

from the start of the 1900s (Benson & Howell 1990, Keast 1995).

E2.2.2 Descriptions

Urban 1: Model Yacht Pond, Centennial Park [AMG 337300E 6247500N] (Fig E2.2)

The freshwater wetlands dominating the sand dunes of the eastern suburbs were

epitomised by two of the four largest wetlands: the Lachlan Swamps (within

Centennial Park, Urban 1) and the Botany Swamps (Urban 2) (Benson & Howell

1990). Although they were modified by water supply schemes, and drained and filled

for industrial and agricultural use, some original sclerophyllous heath, scrub and low

forest of the sand dunes still survives at these sites (Benson & Howell 1990).

Centennial Park is a large (219 hectares) urban parkland that contains 9 lagoons. It

served as Sydney’s main water supply from 1827-1886 (Butlin 1976, Keast 1995), but

now suffers from a heavy influx of road runoff, and a high nutrient burden due to

runoff from the neighbouring agricultural showgrounds, and defecation by a vast bird

population – largely Australian white ibis (Threskiornis molucca), but also a variety

of waterfowl. One small lagoon, Model Yacht Pond (170 m x 80 m), was chosen for

this study as C. longicollis is more abundant here than in the larger lagoons

(Stephenson 1986) where they co-habit with large numbers of E. macquarii (pers.

ob.). It receives road runoff directly through a stormwater inlet. The pond is largely

unshaded, with trees (Casuarina spp.) bordering one bank. There are some submerged

macrophytes and some areas of emergent reeds near the banks. Short-finned eels

(Anguilla australis) are also present (pers. ob.).

After only three C. longicollis had been sampled, Centennial Park & Moore Park

Trust personnel requested that no further tissue samples be taken from captured

turtles, so work at the site ceased.

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Urban 2: Botany Swamps, Eastlakes [AMG 335250E 6244150N] (Fig E2.2)

Dam construction in the late 1800s for increased water storage capacity would have

replaced smaller, impermanent expanses of open water, with a permanently increased

water level and reduced wetland diversity (Benson & Howell 1990). The swamps are

now a chain of ponds that extend for 4.2 km from Gardeners Rd, Eastlakes, to

Foreshore Rd, Botany, with turtles captured north of Wentworth Avenue, above the

weir, in the northernmost lagoon. This section is nestled between The Lakes private

golf course and Eastlakes public golf course and receives wet weather runoff from the

fertilised grounds. The ponds are very open, with disjointed areas of riparian

vegetation, including some reed beds. Submerged macrophytes are also sparse in most

areas. Although the area is solely composed of sandhills, there is a thick layer of black

mud on the bottom of the lakes (Mackay 1949). Originally these swamps were

continuous with the Lachlan Swamps of Centennial Park to the north and the ponds

around Botany in the south, but subsequent land reclamation and developments have

broken the connection (Butlin 1976). European carp (Cyprinus carpio) and long-

finned eels (Anguilla rheinhardtii) are present (pers. ob.).

Urban 3: Sir Joseph Banks Park, Botany [AMG 333150E 6241350N] (Fig E2.2)

Botany is an industrial and residential municipality originally consisting of extensive

sand dunes covered in Eastern Suburbs banksia scrub and draining into a system of

freshwater swamps (Benson & Howell 1990). Industry has long dominated in this

area, with over half of the total liquid factory wastes of Sydney generated in the

Botany Bay drainage basin and 18% of these generated in the old industrial area

around Botany (Butlin 1976). Sir Joseph Banks Park contains a long narrow stretch of

ponds (1300 m long) running along the coast of Botany Bay near the Sydney

International Airport. Three of the variable number of ponds (dependent on rainfall)

were sampled, but not the one pond that receives road runoff. All ponds receive input

of petroleum liquids from aircraft activities at the neighbouring airport (Butlin 1976,

Butlin 1977). The area is planted with native vegetation, although exotics are also

present. Paths run through the area, but the bush is generally thick around the areas

sampled. Pond 1 (35 m x 15 m) is more open than the others, usually with a

cyanobacterial bloom, sparse to dense aquatic weed, clumps of bullrushes (Typha sp.),

and a depth of up to 1.1 m. Pond 2 (25 m x 20 m ) is very swampy with a depth of up

to 1 m and very dense water weed throughout. The surrounding vegetation is very

194

thick although the pond receives a lot of sunlight and heat retention is high due to the

water weed and the shallow depth. The majority of turtles were caught in Pond 2, so

water and sediment samples from this lagoon were used for analyses (Section E2.3.2).

Pond 3 (75 m x 50 m) was densely grown with Typha sp., so nets were placed in

localised open areas. The surrounding bushland is dense, and an elevated boardwalk

crosses the pond.

Urban 4: Bicentennial Park, Rockdale [AMG 328650E 6240250N] (Fig E2.2)

Bicentennial Park contains a single waterbody (500 m x 50 m) with an island lying

within the northern section which is continuous with the southern section although a

footbridge crosses between the two. Only the northern section was sampled. Although

a depth of 4 m is reached, it did not exceed 1.1 m in sampled areas. The pond is

surrounded by open recreational parkland to the west, a major road to the south,

shrubs to the east and some shrubland and housing to the north. Three large drains

empty into the waterbody. Submerged macrophytes are dense at the edges.

Park 1: Lake Toolooma, Heathcote National Park [AMG 314300E 6220500N] (Fig

E2.3)

Lake Toolooma is a single waterbody (500 m x 50 m) fed by a small creek (Coutts

Creek), surrounded by bushland, and created by the presence of a weir at its northern

end. Upstream there is a scout camp (Camp Coutts) which has pit toilets and some

seepage may enter the lake from these. There is a major road (Princes Highway) 1 km

up the hill to the east, separated from the lake by native bushland. The centre of the

lake (> 6 m from the bank) is 2-4 m deep. Spikerushes (Eleocharis sp.) concentrate

within 6 m of the bank and there are submerged macrophytes, and floating

bladderwort.

195

Park 2: Kangaroo Creek, Royal National Park [AMG 320100E 6227750N] (Fig

E2.3)

The world’s second designated national park, the Royal National Park, was

established in 1879, so has remained protected from development and other

potentially polluting activities. Kangaroo Creek is a small stream with good water

quality (Shaun Dwyer pers. comm.) that opens up in its lower reaches to 20 m width.

It flows into the Hacking River 100 m above Audley Weir. The current weir, which

prevents flow in all but heavy wet weather and prevents mixing with tidal waters, was

constructed in 1950, but previous weirs had existed since 1883. Dense native eucalypt

forest and occasional rocky outcrops border the creek, so access to the site was by

rowboat. There are several stands of Typha sp. and Eleocharis sp., with submerged

macrophytes (Vallisneria gigantea and more predominantly the introduced Elodea

canadensis) only dense in the lower reaches towards the weir. The upper limit of

trapping was at the point where the creek narrows and shallows so that passage by

boat becomes impossible. The lower limit was at a large Typha sp. bed about 300m

upstream of the boatshed and upstream of the small creek that runs down Platypus

Gully.

Park 3: Jibbon Lagoon, Royal National Park [AMG 330550E 6226620N] (Fig E2.3)

There are two lagoons south of Jibbon Beach near the small township of Bundeena at

the northern end of the Royal National Park. The northernmost of these was dry

during the sampling period. Turtles were captured in the southern lagoon (150 m x

100 m) which was shallow (0.7m), swampy and largely grown out with Eleocharis sp.

with a corticate Charalles alga the primary submerged aquatic plant. The lagoon is

nestled in a gentle depression and is surrounded by native forest on the northern side

and low banksia heathland to the south, and is open to sunlight. This site is susceptible

to sea spray (ocean 200 m distant).

Park 4: Marley Lagoon, Royal National Park [AMG 328210E 6223730N] (Fig E2.3)

Marley Lagoon (300 m x 200 m) lies in a gentle depression 4.5 km walk south of the

small township of Bundeena. It has a sandy substrate and is bordered by large sand

dunes to the east and low native coastal heathland in the other directions with the

taller Leptospermum sp. & Acacia sp. growing in the more wind protected areas

around the lagoon. It is very open to sunlight. The water is extremely clear and

197

variably brackish due to its susceptibility to sea spray and the presence of a channel

that may connect the lagoon to the ocean (200 m distant) during high seas. The

shallow (< 1 m) areas are largely grown out with Eleocharis sp., with other emergent

species including Schoenoplectus mucronatus also lining the banks. The water reaches

3m depth in a shallow central bowl, the surface of which is blanketed by a

filamentous alga (Chara sp.). The lagoon is fed by a small creek on the western side.

Sewage 1: Sludge Lagoon, Castle Hill Sewage Treatment Plant

A concrete-lined pit filled with raw sewage and rain water was also sampled (Section

D2.2.5 CH.sludge).

E2.3 Field Work

E2.3.1 Trapping

Trapping effort and method varied between sites. Trapping was initiated in spring

when turtle captures increase and annual reproduction begins. Trapping commenced

on 19 October 2000, and ceased on 24 February 2001. Yabby traps (Section B.1.1.2)

were initially used for trapping turtles at all sites except Lake Toolooma from where a

platypus had been reported (Royal National Park Service pers. comm.). Platypuses

may drown if they follow macroinvertebrate prey into the trap (Serena 1996, Tanya

Rankin pers. comm.), so at this site fyke nets (Section B.1.1.1) were used. Due to

initial low captures, fyke nets were also used at Bicentennial Park, Kangaroo Creek

and Sir Joseph Banks Park. Snorkelling was not possible at urban sites due to poor

water quality or at Jibbons Lagoon due to its shallow swampy nature, but it was

attempted at Lake Toolooma (without success due to poor visibility from turbidity and

depth), Kangaroo Creek and Marley lagoon (both with some success). A summary of

trapping effort is presented (Table E2.1). Only three C. longicollis were captured at

Model Yacht Pond before work was discontinued due to concerns of Centennial Park

& Moore Park Trust personel over tissue sampling.

E2.3.2 Water and Sediment Sampling

At each lagoon two 200 mm depth 50 ml water samples and two 50 ml surface

sediment samples were collected in 70 ml capped sample containers (clean room

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manufactured, Techno-plas, 8 Benjamin St, St Marys, SA 5042) where water depth

was 300-600 mm in order to determine metal levels in the inorganic environment.

Site Month Yabby Effort (h)

Fyke Effort (x

24h)

Snorkelling Effort (h)

U1 (Model Yacht Pond) October 2000 30 - -

U2

(Botany Swamps) October 2000

December 2000 81 66

- -

- -

U3

(Sir Joseph Banks Park)

October 2000 January 2001

42 32

- 16

- -

U4

(Bicentennial Park) October 2000 January 2001

36 30

- 21

- -

P1

(Lake Toolooma) November 2000 February 2001

- -

6 48

- 6

P2

(Kangaroo Creek) November 2000 December 2000

30 84

6 -

- 14

P3

(Jibbon Lagoon) October 2000 77 - -

P4

(Marley Lagoon) October 2000 February 2001

70 70

- -

- 15

Table E2.1 Effort for the three capture methods used at four urban (U1-4) sites and 4

national park (P1-4) sites.

Aquatic physicochemical parameters influence the physical and chemical state of

metals and thus affect metal uptake and toxicity (Beijer & Jernelöv 1979b, Jensen &

Bro-Rasmussen 1992). Changes in metal toxicokinetics within reptiles associated with

changes to water physicochemistry have not been identified (Linder & Grillitsch

2000), so variations in these parameters (pH, dissolved oxygen, salinity, turbidity)

over study sites were assessed, thus helping to identify potential factors for further

study. Data was collected at three urban and three national park sites using a Horiba

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water quality monitor (not available during trips to Centennial Park or Jibbons

Lagoon). A one off reading of water quality at Bicentennial Park was obtained from

Rockdale City Council (pers. comm.), and readings from sites around Sir Joseph

Banks Park and Botany Swamps were obtained from Botany Bay City Council (1999,

pers. comm.).

E2.3.3 Field Turtle Processing

While on site, captured turtles were weighed and the straight carapace length (CL)

measured (Section B.2.2). Blood (Section B.2.5.2) and carapacial bone (Section

B.2.5.1) samples were placed in acid-washed (Section E2.6.1) eppendorf tubes. To

determine if tissue metal concentrations varied with turtle size (CL), carapacial bone

from two urban sites with a large size range of turtles was sampled: C. longicollis (n =

23, CL = 126-210) from Sir Joseph Banks Park and E. macquarii (n = 18, CL = 133-

254) from Bicentennial Park. The influence of sex on metal concentrations in

carapacial bone was examined at two urban sites with high capture rates: Sir Joseph

Banks Park for C. longicollis (5 males, 7 females) and Botany Swamps for E.

macquarii. Botany Swamps had several gravid E. macquarii, so carapace could be

sampled from non-gravid females (n = 5), gravid females (n = 5), and males (n = 5),

but one gravid sample was destroyed during digestion.

Blood volume taken varied, but at least 500 μl of blood was taken (Section B.2.5.2)

from 14 C. longicollis at Sir Joseph Banks (Urban 3) and 14 E. macquarii at

Bicentennial Park (Urban 4) in the adult size range. Turtles were returned to the

water, except for gravid females (determined by palpation; Cagle 1950) which were

removed to the laboratory for induction of egg-laying. Gravid females were also

collected from two Immune Study sites (Section D2.4).

E2.4 Laboratory Turtle Processing

Six C. longicollis that had drowned during the Immune Study (Section D3.1) were

partially thawed at room temperature for dissection. Three were from QH.STP.up

4/1/00 (Dead A, Dead B, Dead F), one from QH.STP.down 5/1/00 (Dead C), one

from RH.STP.down 22/3/00 (Dead D), and one from RH.CONT.down 23/3/00 (Dead

200

E). A hacksaw (300 mm 32T blade; Stanley, Australia) was used to remove the head

and limbs and saw through the bridge and the pelvic girdles where they join the

plastron. A stainless steel surgical scalpel was used to cut through the peritoneum.

Whole liver, whole kidneys (which lie against the inside of the carapace), femur,

claws (removed at the skin’s edge with a scalpel) and carapacial bone (Section

B.2.5.1) were removed and frozen for later analysis.

E2.4.1 Hard Tissue Structure

A seventh turtle (Dead G), found drowned trapped against an effluent outfall grate at

RH.STP.down in January 2000, was removed and frozen. Carapace was sampled

latitudinally and longitudinally from Dead G and femur sampled from Dead E for

microscopic examination of bone structure in order to guage the extent of tissue

remodelling (i.e. bone deposition and re-mobilisation), and to determine if a laminated

structure is present (which may indicate time periods of metal deposition). Samples

were oven-dried at 40ºC and then sectioned using a geological band saw (Accutom-2,

made by Struers) using a 0.38 mm thick diamond blade (Leco Australia) at 700-800

revolutions/min, and embedded in a cylindrical mount of resin 25 mm diameter and

10 mm high. Polished to a 0.25-0.5 micron finish.

Mounts were examined using light microscopy, and the carapacial bone mounts were

also carbon-coated for backscattered and X-ray emission scanning electron

microscopy (SEM). Secondary Ion Mass Spectrometry (SIMS) has been successfully

used to map changes in Pb concentration over layers in crocodile osteoderms

(Twining et al. 1999), so C. longicollis carapacial bone was examined for lamination

to ascertain if SIMS could be used to detect changes in metal deposition over

laminations and, possibly, therefore over time, thus allowing turtles to be used as

records of historical aquatic metal loads.

E2.5 Preparation for Metal Analysis

E2.5.1 Tissues

Tissue samples from field and laboratory turtles were cleaned if necessary, weighed,

dried to constant weight in a 40 ºC or 60 ºC oven, re-weighed and then microwave-

digested prior to analysis by ICP-MS and ICP-AES. Glassware used in this process

201

(and eppendorf tubes used for blood collection) was previously soaked overnight in

5% Decon, rinsed twice in distilled water, then soaked overnight in 5% nitric acid and

rinsed twice in deionised water (reverse osmosis, Ionpure) before being drained

upside down on an absorbent sheet in a closed cupboard. Swann-Morton, #0208,

carbon steel, gamma-sterilised surgical blades were used on scalpels. EDTA

(diSodium ethylenediamine-tetraacetate, MWt 372.24, M&B Laboratory Chemicals,

May & Baker Ltd, Dagenham, England, #5246/18/67) solutions were made up in

deionised water. Pipette tips were from Trace Plastics, Biosciences Pty Ltd, Australia.

Carapace

Scute and germinative layers were removed from the bone with a scalpel and the

remaining carapacial bone weighed to 0.005 mg. To remove loosely bound

contaminants, samples were then briefly rinsed in deionised water, agitated for 10 sec

in 10% HCl (sub-boiled ultra-pure), briefly rinsed in deionised water, agitated in 100

mM EDTA for 1 min, then finally rinsed 3 times in deionised water before drying in

glass beakers or petri dishes.

Blood

Tubes containing blood samples were weighed, dried at 40 °C (usually) or 60 °C

(occasionally) dependent on other oven users, then re-weighed. An attempt was made

to resuspend dried blood by sonication (power output of 30-40 watts) in 0.5% triton

X-100, but this was not successful, and dried blood was simply scraped into digestion

vessels using a 200 μl Gilson pipette tip. The tube was then rinsed twice with 500 μl

of 50% HNO3 using a 1000 μl Gilson pipette tip with the end cut off, and rinses added

to the digestion vessel. The empty eppendorf tubes were then re-weighed after air

drying, and this weight subtracted from weights of tubes with both wet or dry blood.

Egg (Section F)

Only one egg per clutch was analysed as within-clutch variations in metal

concentration are low (< 0.07%, Stoneburner et al. 1980; Sakai et al. 1995). Eggs

were thawed at room temperature, then rinsed in deionised water, rinsed in 100 mM

EDTA, rinsed 3x in deionised water, blotted dry with a tissue, then weighed. They

were cut with a scalpel around the equator and the contents emptied into a glass

202

beaker, weighed, and thoroughly mixed before drying. Eggshells were washed in

deionised water before drying.

Liver and Kidney

Not all metals distribute evenly along the length of internal organs (Woodling et al.

2001), so the whole liver and each kidney were placed into glass beakers, weighed,

roughly homogenised with a scalpel, and the liver then halved for duplicate samples.

Femur

One femur from each turtle was thawed and any remaining flesh or cartilage removed

with a scalpel, including scraping with the blunt edge. Femurs were weighed to three

decimal places, agitated for 60 s in 100 mM EDTA, agitated in deionised water for 60

s, then given one further 10 s rinse in deionised water. Femurs were then placed on

individual glass petri dishes in a 40 ºC drying oven.

Claw

All four claws were removed from each of the four feet at their base (skin’s edge)

with a pair of scissors, leaving the ball behind but taking the socket of the joint. Claws

were weighed, rinsed twice in deionised water, soaked for one hour in 100 mM

EDTA, rinsed twice in deionised water, rinsed once in 10% HCl (10 s), then rinsed

twice in deionised water and placed in a 60 ºC drying oven.

E2.5.2 Environmental Samples

Sediment

Sediment samples were freeze-dried (-20°C, 0.100mbar, Christ Gamma 2-20) and

ground with an acetone-washed mortar and pestle. Any large organic particles (e.g.

leaf parts, roots) were picked out after freeze-drying.

Water

Following the standard ANSTO method (Henri Wong pers. comm.), water samples

(50 ml) were acidified with 0.25 ml of sub-boiled ultrapure 70% HNO3, then filtered

into 50 ml plastic tubes using a 20 ml syringe (Terumo Medical Corporation, Elkton,

MD, USA) which had rinsed with Milli Q water, and a sterile 0.45 μ filter unit

203

(Millipore S.A., Molsheim, France). Long glass tubes with ball lids were cleaned

(Section E2.5.1), and 20 ml of water sample added. To this was added 2 ml sub-boiled

HNO3 and 200 μl sub-boiled HCl. Tubes were placed in a block digester and the

temperature ramped at 2ºC/min to a max of 90ºC and held there for 3 h. Once cooled,

the water was poured into 50 ml tubes.

E2.6 Metal Analysis

E2.6.1 Quality Control

One to three blanks were analysed every ICP-MS/AES run (every 10-40 samples).

Blank values were subtracted from metal concentrations found in experimental

samples (μg/L) before conversion of concentrations to μg/kg of original solid sample.

Appropriate certified reference standards were analysed every ICP-MS/AES run

(every 10-40 samples). Dogfish muscle (standard reference material DORM-2;

National Research Council Canada, Ottawa, Ontario) was run with liver, kidney,

blood and egg contents samples. Oyster (standard reference materical 1566a) and/or

bone ash (standard reference materical 1400), both from the National Institute of

Standards & Technology (Gaithersburg, MD, USA) were run with carapace, femur

and eggshell samples. Standards were dried and digested in the same manner as

experimental samples.

To gauge precision, fourteen carapace samples, between the weights of 20 and 200

mg, from one turtle (Dead E) were digested and analysed in three separate ICP-

MS/AES runs.

E2.6.2 Microwave Digestion of Samples

Teflon digestion vessels (100 ml capacity) were soaked overnight in 5% Decon then

rinsed twice in deionised water. Vessels were subsequently boiled for 1 h in 50% sub-

boiled ultrapure HCl, cooled (room temperature for 1 h), then boiled for 1 h in 50%

sub-boiled nitric acid (analytical reagent; equal volumes 70% HNO3 and water) and

again left to cool. Vessels were then twice rinsed in deionised water and left upside

down on paper towel and covered to air dry overnight.

204

Approximately 200 mg of dried tissue was weighed into digestion vessels, on an

A.N.D. HA-202M analytical balance. The exceptions were blood, where

approximately 100 mg of dried tissue was weighed, and some smaller carapace

samples where a minimum of 60 mg was weighed. Three ml H2O and 3 ml 30% H2O2

(either Riedel-deHaen #18312 extra pure, or Univar Analytical Reagent #260-2.5L PL

which contains up to 0.001% tin) were then added to each vessel. Lids were placed

lightly atop vessels and left overnight, to allow escape of CO2 from organic material.

The following day, 3 ml concentrated HNO3 (ultrapure sub-boiled) was added and the

vessel + lid + pre-digest weighed. For sediment samples, 250μl 40%w/w hydrofluoric

acid (Univar Analytical Reagent; Ajax Chemicals, 9 Short St, Auburn NSW 2144,

Australia) was also added. Vessels were placed in the carousel and adjustable screws

tightened. The carousel was then placed in the microwave digester (Milestone mls

1200 mega; Via Fatebenefratelli 1/5, 24010 Sorisole, Italy) and run for 5 min at

250W, 5 min at 400W & 15 min at 500W. After 1 h the carousel was removed to the

fume hood and left to cool for 1 h.

Vessel + lid + digest was then weighed and the digest discarded if loss exceeded 10%.

Digests were poured into pre-weighed labelled 50 ml tubes, the vessel rinsed twice

with deionised water and the rinses added and the final volume made up to

approximately 25 ml with deionised water. The tube + digest was then weighed and

the digest weight calculated by subtraction of the tube weight.

E2.6.3 ICPMS & ICPAES Digest Analysis

Digested tissue samples were analysed for 40 different metals (Table E2.2) using

Inductively Coupled Plasma Mass Spectrometry (ICP-MS) and Inductively Coupled

Plasma Atomic Emission Spectrometry (ICP-AES) using ANSTO methods VEC-I-9-

03-007 and VEC-I-9-03-002 (Henri Wong pers. comm.) respectively (conducted by

the Chemistry Department, Environmental Science Division, ANSTO). Detection

limits were 1μg/kg for all elements except Al, Bi, Fe, In, Sc (10μg/kg), Ca, Y

(30μg/kg), and K (50μg/kg).

205

Atomic

# Mass

Element

3 6.9 Li Lithium

4 9.0 Be Beryllium

11 22.9 Na Sodium

12 24.3 Mg Magnesium

13 26.9 Al Aluminium

19 39.0 K Potassium

20 40.0 Ca Calcium

21 44.9 Sc Scandium

22 47.8 Ti Titanium

23 50.9 V Vanadium

24 51.9 Cr Chromium

25 54.9 Mn Manganese

26 55.8 Fe Iron

27 58.9 Co Cobalt

28 58.6 Ni Nickel

29 63.5 Cu Copper

30 65.3 Zn Zinc

33 74.9 As Arsenic

34 78.9 Se Selenium

38 87.6 Sr Strontium

39 88.9 Y Yttrium

40 91.2 Zr Zirconium

42 95.9 Mo Molybdenum

47 107 Ag Silver

48 112 Cd Cadmium

49 114 In Indium

50 118 Sn Tin

51 121 Sb Antimony

52 127 Te Tellurium

55 132 Cs Cesium

56 137 Ba Barium

57 138 La Lanthanum

71 174 Lu Lutetium

74 183 W Tungsten

80 200 Hg Mercury

81 204 Tl Thallium

82 207 Pb Lead

83 208 Bi Bismuth

90 232 Th Thorium

92 238 U Uranium

Table E2.2 The atomic number, atomic mass, symbol and name of the forty elemental

metals analysed by ICP-MS (roman type) and ICP-AES (italics).

206

CHAPTER E3: METAL BIOACCUMULATION RESULTS

E3.1 Turtle Captures

Three turtle species were captured (Table E3.1), with only Chelodina longicollis captured at

all urban (n = 4) and all national park (n = 4) sites. Emydura macquarii was captured at three

urban (not Model Yacht Pond) and two national park (Kangaroo Creek and Lake Toolooma)

sites. Elseya latisternum was limited to two urban sites: Botany Swamps and Bicentennial

Park. Twelve of the 75 C. longicollis from the sewage sludge lagoon (Section D2.2.5) were

also included in this study to represent a highly polluted environment.

Site C. longicollis E. macquarii El. latisternum

U1 (Model Yacht Pond) 3 - -

U2 (Botany Swamps) 2 22 2

U3 (Sir Joseph Banks Park) 36 1 -

U4 (Bicentennial Park) 8 22 1

P1 (Lake Toolooma) 5 3 -

P2 (Kangaroo Creek) 2 9 -

P3 (Jibbons Lagoon) 9 - -

P4 (Marley Lagoon) 10 - -

Sludge Lagoon 12 - -

Table E3.1 The number of turtles captured at the four urban (U1-U4) and four national park

(P1-P4) sites, and the number of turtles sampled from the sludge lagoon. Trapping effort was

not equal at all sites (Table E2.1).

E3.1.1 Turtle Size and Body Condition

Carapace length (CL) of C. longicollis varied significantly among sites (ANOVA: F = 2.413,

df = 8, 61, p = 0.025), although Bonferroni-adjusted pairwise comparisons did not identify

any pair of samples as significantly different. Carapace length of E. macquarii also varied

significantly among sites (excluding the single Sir Joseph Banks Park E. macquarii, F =

7.323, df = 3, 40, p = 0.001), with the Kangaroo Creek sample being significantly smaller

than the other three samples (Bicentennial Park, Botany Swamps, Lake Toolooma).

Kangaroo Creek was the only site with no E. macquarii of CL > 190 mm.

207

Using the indices generated from the Sydney Survey (Section C3.4.4), there was no

significant difference in turtle body condition between sites for either C. longicollis

(ANOVA: F = 1.302, df = 8, 61, p = 0.260), or E. macquarii (ANOVA: F = 2.225, df = 3, 40,

p = 0.100). As a check that these indices, derived from other Sydney samples, did not vary

with size (CL), regression of the body condition indices for each species against CL was

undertaken; in neither species was there a significant relationship with CL (C. longicollis: r2

= 0.020, p = 0.246; E. macquarii: r2 = 0.051, p = 0.134).

E3.2 Metal Analysis

E3.2.1 Quality Control

Accuracy of metal detection (Section E2.6.1) varied over metals and certified reference

standards. For DORM-2 (dogfish muscle), detection was never higher than the certified

reference value. Metal concentrations were usually within 15% of the certified value (As, Hg,

Al, Cu, Se, Zn), although some were only reliably within 50% (Ni, Mn, Fe), and Cr was only

within 60%. A minimum of 250 mg tissue must be analysed for values to be certified and

where this was done (n = 4) Fe and Mn were detected within 25% of reference values, and Ni

within 40%. For oyster, detection was never higher than the certified reference value. Metal

concentrations were usually within 10% (V, Cd, Mn, Cu, Zn, Na, K) to 20% (Ni, As, Se, Fe,

Mg, Ca, Sr) of the certified value, although Al was only consistently within 40%. For bone

ash, Mn was over-estimated by an average of 15%. Other metals were not over-estimated and

were within 5% (Pb, Ca, Sr, Na, K) or 10% (Cu, Fe, Zn, Mg) of the certified value, except for

Al which was only within 25%.

In summary, metal concentrations were generally underestimated but were within 0-20% of

certified values. Exceptions are Mn, which was overestimated by 15% in bone, and Al which

is underestimated by up to 40%. Detection of Mn, Ni, Cr and Fe is as low as 50% in

proteinaceous tissues. Also, although the given Ti concentration in bone ash was not

certified, detection of Ti in bone ash was highly variable (80 ± 40 mg/kg S.E., range 16-136

mg/kg, n = 6).

For the fourteen carapace samples (20-200 mg) from turtle ‘Dead E’ (Section E2.6.1),

variation was mostly within 30% for the eight metals that were always detected (Ti, Al, Mn,

208

H He

Li Be B C N O F Ne

Na Mg Al Si P S Cl Ar

K Ca Sc Ti V Cr Mn Fe Co Ni Cu Zn Ga Ge As Se Br Kr

Rb Sr Y Zr Nb Mo Tc Ru Rh Pd Ag Cd In Sn Sb Te I Xe

Cs Ba La Hf Ta W Re Os Ir Pt Au Hg Tl Pb Bi Po At Rn

Fr Ra Ac Unq Unp Unh

Lanthanides Ce Pr Nd Pm Sm Eu Gd Tb Dy Ho Er Tm Yb Lu

Actinides Th Pa U Np Pu Am Cm Bk Cf Es Fm Md No Lr

not detected environment only turtles only environment & turtles

Figure E3.1 Periodic table showing the 40 metals analysed using ICPMS or ICPAES in bold, and the detection of these

elements in environmental and turtle tissue samples.

209

Zn, Mg, Ca, Sr, Ba), except where dry weight of carapace was less than 60 mg (n = 6) when

lower concentrations (as low as 50% of other values) were sometimes obtained. For this

reason 60 mg was set as the minimum dry weight for sample analysis.

E3.2.2 Metal Detection

Of the 40 metals analysed, 17 were detected in both environmental and turtle tissue samples,

12 were detected in the environment but not tissues, and only Se was detected in tissues but

not the environment (Figure E3.1). Ten metals (Sc, Y, Ag, In, Te, W, Hg, Tl, Bi, Lu) were

not detected in any environmental or turtle sample, and these are excluded from further

discussion. In lists of metals, non-essential (Pb, Al, Ti, Zr, Ni, Cd, As, Sn, V, Sb), essential

(Cr, Mn, Co, Cu, Se, Mo, Fe, Zn), alkaline earth (Be, Mg, Ca, Sr, Ba), and alkali metals (Li,

Na, K, Cs), and the lanthanides/actinides (La, Th, U) are separated by ‘/’.

E3.3 Water and Sediment

E3.3.1 Water Quality

Physicochemical water quality measurements varied over the six lagoons (Table E3.2). Sir

Joseph Banks Park and Bicentennial Park have a pH > 8, and Lake Toolooma and Kangaroo

Creek have a pH < 6.5. All the lagoons have higher salinity than expected for freshwater

lakes, but only Marley Lagoon is above the range for natural freshwaters (ANZECC &

ARMCANZ 2000a). All sites have oxygen concentrations higher than 5 mg/L.

Site

Temperature

(°C)

pH Salinity

(μS/cm)

Dissolved O2

(mg/L)

U2 (Botany Swamps) 24.7 7.3 276 6.83

U3 (Sir Joseph Banks Park) 27.1 9.0 878 7.71

U4 (Bicentennial Park) 28.2 8.3 455 9.87

P1 (Lake Toolooma) 24.7 6.1 130 7.06

P2 (Kangaroo Creek) 22.9 5.7 144 7.91

P4 (Marley Lagoon) 29.9 6.7 4855 5.98

Table E3.2 Physicochemical water quality measurements at three urban (U2, U3, U4) and

three national park (P1, P2, P4) sites.

210

E3.3.2 Metals in Water

Twelve metals were detected in water from at least one lagoon (Table E3.3). Eighteen metals

were not detected in water (Ti, Al, Cd, Ni, As, Sn, Zr, Sb, V / Cr, Co, Se, Mo / Be / Cs / La,

Th, U). Centennial Park (U1) and Botany Swamps (U2) have water concentrations of Pb, Cu,

Zn and Fe above the trigger value. Lake Toolooma (P1) and Jibbons Lagoon have water Zn

concentrations above the trigger value. The concentration of Ca dictates the hardness of

water, with all the waters soft (0-24 mg Ca/L), except for P4 which is moderate (25-47 mg

Ca/L), and U1 and U2 which are extremely hard (> 160 mg Ca/L) (ANZECC & ARMCANZ

2000b).

Individually

U1, U2, U3, U4

Urban

+/-SE

Individually

P1, P2, P3, P4

Park

+/-SE

trigger

value*

Non-essential metals

Pb 855, 721, nd, nd 394+/-229 nd, nd, nd, nd nd 3.4

Essential metals Mn 91, 190, nd, nd 70+/-45 nd, nd, nd, nd nd 1700 Cu 15, 7, nd, nd 5+/-4 nd, nd, nd, nd nd 1.4 Zn 57, 26, 8, 6 24+/-12 10, 5, 16, 4 9+/-3 8 Fe 905, 1810,

nd, nd 679+/-433 nd, nd, nd, nd nd 300

Alkaline earth metals

Mg 2746, 3583, 6031, 6030

4598 +/-845

1768, 1552, 5315, 86923

23890 +/-21029

-

Ca 224544, 315347, 22159, 21131

145795 +/-74036

3041, 1146, 1659, 28321

8542 +/-6605

-

Sr 62, 103, 160, 114 110+/-20 15, 12, 32, 534 148+/-129 - Ba 37, 47, 13, 42 35+/-8 6, 7, 8, 33 13+/-6 -

Alkali metals

Li nd, nd, 13, 11 6+/-3 nd, nd, nd, 6 2+/-2 - Na 13507, 13139,

28653, 30360 21415

+/-4685 11462, 10239, 36507,

752829 202759

+/-183457 -

K 1711, 2981, 3365, 5540

3399 +/-796

295, 479, 3147, 28216 8034 +/-6759

-

Table E3.3 Metal concentrations (μg/L) in water samples from four urban (U1-U4) and four

national park (P1-P4) sites in grey, with mean +/- standard error in black. nd = not detected.

*Aquatic metal concentrations above the trigger value (μg/L) could be of concern for wildlife

(ANZECC & ARMCANZ 2000b). U1-U4 and P1-P4 defined in Table E2.1.

211

There are no significant differences between urban (n = 4) and park (n = 4) sites (Mann-

Whitney U-tests) in the water concentrations of 11 metals (Pb / Mn, Cu, Zn, Fe / Mg, Ca, Sr /

Li, Na, K). Water Ba concentration is 2.6x higher in urban than park sites (Mann-Whitney U-

test statistic = 15.0, p = 0.043).

E3.3.3 Metals in Sediment

Twenty-nine metals were found in sediment from at least one lagoon (Table E3.4). Only Se

was not detected in any sediments. For no metal are the four highest concentrations all within

urban sites or all within park sites. Excluding Zr and La, for which data is incomplete, three

of the four highest concentrations are within park sites for no metals, and are within urban

sites for 15 metals (V, Ni, Pb, As, Sb / Mn, Mo, Cu, Zn, Fe / Ca, Sr, Ba / Li, Cs). One urban

(U3) and two park (P2, P4) sites do not commonly have the higher concentrations of metals.

There were no significant differences between urban and park sites in the concentrations of

any sediment metal (Zr and La excluded, Mann-Whitney U-tests, 4 urban and 4 park sites).

E3.3.4 Metals in Water, Sediment, and Carapace

There are no significant correlations between metal concentrations in water, sediment and

mean carapace value across the eight sites for six metals (Fe / Mg, Ca, Sr, Ba / K) (Tables

E3.10-E3.17). For Pb, there is a significant correlation (Pearson’s p = 0.922; Bonferroni

adjusted probability = 0.003) between concentrations in water and carapace, but no

correlation between water and sediment or sediment and carapace. Significant correlations

were detected between water and sediment concentrations but not carapace vs water or

sediment for Cu (Pearson’s p = 0.974; Bonferroni adjusted probability < 0.001), and Zn

(Pearson’s p = 0.958; Bonferroni adjusted probability = 0.001). For Mn the correlation

between sediment and water concentrations approaches significance (Pearson’s p = 0.791;

Bonferroni adjusted probability = 0.058). Ti, Al, Zr and Cr were not detected in water, and

concentrations were not significantly correlated between sediment and carapace.

There was no effect of the concentration of Ca in water (water hardness) on the metal

concentrations in carapace for most metals; however, there was an effect of Ca concentration

in water on carapace Se (r2 = 0.627, F = 10.096, df = 1, 6, p = 0.019) and carapace Pb (r2 =

0.675, F = 12.483, df = 1, 6, P = 0.012). In both cases the relationship was positive.

212

U1 U2 U3 U4 P1 P2 P3 P4 Non-essential metals (μg/kg)

Be nd nd nd nd nd nd 2610 nd V 20547 16991 8128 19961 27023 7100 16228 12972 Ni 7037 6996 1392 6875 4352 1424 6477 2390 Pb 219415 155942 4438 32284 31744 7337 736776 10187 As 1580 2555 147 2157 4727 339 159 1521 Zr na na 6736 17860 111290 na 8080 10345 Cd nd nd nd 638 nd nd 917 nd Sn 7103 4827 nd 1243 1472 nd 86810 357 Sb 2699 1886 nd 913 384 nd 21077 94 Cs 775 528 nd 446 nd nd 1583 149

Essential metals (μg/kg)

Cr 18062 13167 5546 17214 18937 5403 11973 8846 Mn 67342 84295 15123 64120 25138 10754 35271 17390 Mo 2643 nd 185 3674 nd nd 61 nd Co 2234 2034 611 3250 2811 613 2120 931 Cu 59822 26673 368 11457 9536 817 9271 1418

Essential (Zn, Fe) and non-essential (Al, Ti) metals (mg/kg)

Zn 366 212 8 100 59 9 135 10 Fe 5487 6755 2666 8301 8609 2469 3867 3292 Al 16681 15430 6964 11651 23960 7630 16126 12068 Ti 1105 1070 392 710 3262 660 814 590

Alkaline earth metals (mg/kg) Mg 1385 1186 15 67 109 207 43 16 Ca 3396 2216 524 7923 782 161 3899 1038 Sr 37 35 13 46 10 18 42 30 Ba 98 134 76 93 39 34 217 48

Alkali Metals (mg/kg)

Li 6.9 4.7 6.7 19.8 6.6 1.1 4.8 5.6 Na 1058 1020 497 915 426 130 2058 3384 K 3689 3137 3219 3843 4570 1473 4119 1946

Lanthanides & actinides (μg/kg)

La na na 3409 2771 nd na 28445 6964 Th 2240 3086 1114 1471 1310 2093 2513 2553 U 418 453 590 1356 795 nd 545 267

Table E3.4 Metal concentrations in sediment samples from four urban (U1-U4) and four

national park (P1-P4) sites. The four sediments with the highest concentration of metal are

shown in bold. na = not analysed, nd = not detected.

213

E3.4 Organ Metal Distribution

Fifteen metals (Pb, Ti, Ni / Cr, Mn, Cu, Se, Zn, Fe / Mg, Ca, Sr, Ba / Na, K) were detected in

at least one tissue (liver, kidneys, femur, claws, carapace) of C. longicollis (n = 6) (Table

E3.5) collected from creeks in north-western Sydney (Section D2.4). Three more metals were

present only at or near detection limits (Al & Zr in carapace, Mo in liver & kidney), and 12

metals (As, Cd, Sn, V, Sb / Co / Be / Li, Cs / La, Th, U) were below detection limits in all

tissues.

Correlations between metal concentrations among tissues are only explored where more than

four non-zero values are available for a tissue. There are no significant correlations in Ti, Pb,

Se, Cu, Cr, Mg, Ca, or Na concentrations among the five tissues. For Zn concentrations there

is a significant correlation (Pearson’s p = 0.967, p = 0.010) between carapace and claw, but

not for other pairs of tissues. For Fe concentrations there is a significant correlation

(Pearson’s p = 0.929, p = 0.007) between liver and claw, but not between other pairs of liver,

claw and kidney. For Ba there are significant correlations between concentrations in carapace

and claw (Pearson’s p = 0.945, p = 0.026), carapace and femur (Pearson’s p = 0.996, p <

0.001) and claw and femur (Pearson’s p = 0.951, p = 0.022), but not between liver or kidney

and the other three tissues. For Sr there are significant correlations between most pairs of

liver, carapace, claw, and femur (carapace vs liver Pearson’s p = 0.963, p = 0.012, femur vs

liver Pearson’s p = 0.949, p = 0.023, claw vs carapace Pearson’s p = 0.927, p = 0.047, femur

vs carapace Pearson’s p = 0.997, p < 0.001, correlations between claw and liver, and claw

and femur are just below significance), but there are no correlations between concentrations

in kidney and other tissues. For Mn there is a significant correlation between femur and

carapace (Pearson’s p = 0.935, p = 0.037), and between kidney and liver (Pearson’s p =

0.980, p = 0.003), but not between other pairs of tissues. For K, there is a significant

correlation between femur and claw (Pearson’s p = 0.926, p = 0.048), but not between other

pairs of tissues.

Although Sr, Ba, and the other alkaline earth metals, are thought to distribute throughout the

body in proportion to Ca, with predominant deposition in bony tissues (Camner et al. 1979,

Ware 1988a), in this study there are no significant correlations between either Ba or Sr and

Ca concentrations in carapace, femur or claw, or between Ba and Ca in liver, or Sr and Ca in

kidney. However, there are significant correlations between Sr and Ca in liver (Pearson’s p =

0.973, p = 0.001) and between Ba and Ca in kidney (Pearson’s p = 0.924, p = 0.028).

214

Liver Kidney Carapace Femur Claw Lead range 0-2086 0-1358 0-24018 0-9857 0-9435 mean 573.3 - 11804.7 5055.5 2293.5 SD 846.4 - 9006.9 5000.8 2903.1 n 6(3) 5(4) 6(4) 6(2) 6(4)

Titanium range 845-2128 864-2154 20304-69467 - 0-6829 mean 1402.5 1457.6 53340.8 - - SD 532.5 511.7 21242.5 - - n 6(0) 5(0) 6(0) 6(6) 6(5)

Nickel range - - - 413-750 - mean - - - 612.0 - SD - - - 126.2 - n 6(6) 5(5) 6(6) 6(0) 6(6)

Chromium range 0-899 0-959 0-5125 - - mean 307.5 191.8 1872.5 - - SD 371.2 428.9 1989.1 - - n 6(3) 5(1) 6(1) 6(6) 6(6)

Manganese a a b b range 3884-24337 1283-15235 12838-27138 6740-20763 0-5622 mean 9374.3 5207.2 19617.0 15775.8 2932.3 SD 7674.1 5743.4 5151.1 5226.6 2407.2 n 6(0) 5(0) 6(0) 6(0) 6(2)

Copper range 17998-49667 4491-10282 0-355 - - mean 37262.8 6994.6 - - - SD 13366.2 2149.8 - - - n 6(0) 5(0) 6(5) 6(6) 6(6)

Selenium range 2538-5575 2794-5016 - 1238-2502 0-12985 mean 4195 4067.2 - 2082.3 6162.5 SD 1089.9 1074.1 - 458.9 4412.1 n 6(0) 5(0) 6(6) 6(0) 6(1) Table E3.5 Summary statistics for metal concentrations (μg/kg tissue) in tissues from six turtles. Value in parentheses under ‘n’ is number of individuals with metal concentrations measured at 0. Means and standard deviations are provided only for tissues where more than one value above 0 was detected; means and standard deviations include the 0 values. Significant correlations between tissues for each metal are indicated by corresponding letters.

215

Liver Kidney Carapace Femur Claw Zinc a a range 42-108 38-76 67-125 47-180 190-370 mean 74.2 59.4 96.3 118 281 SD 24.0 16.3 31.0 46.8 76.0 n 6(0) 5(0) 6(0) 6(0) 6(0)

Iron a a range 741-3137 301-450 - - 18-134 mean 1630.8 346.8 - - 46.0 SD 908.1 60.4 - - 45.3 n 6(0) 5(0) 6(6) 6(6) 6(0)

Magnesium range 361-692 434-669 2340-3001 1042-2426 390-787 mean 475.5 575.0 2761.2 1998.2 610.7 SD 129.1 92.0 240.7 492.7 158.4 n 6(0) 5(0) 6(0) 6(0) 6(0)

Calcium range 934-3396 1178-2652 199351-

259409 121261-191386

19755-62717

mean 1782.7 1724.0 233333.5 164800.0 45785.8 SD 952.8 560.3 24201.6 23737.2 15102.0 n 6(0) 5(0) 6(0) 6(0) 6(0)

Strontium a,b a,c,d b,c d range 0.95-7.34 1.59-3.55 229-892 168-707 37-171 mean 3.027 2.147 389.2 299.4 7.8 SD 2.285 0.807 249.8 203.8 48.2 n 6(0) 5(0) 6(0) 6(0) 6(0)

Barium a,b b,c a,c range 0-3.17 0-1.06 80 -426 54.1-362.4 5.6-93.1 mean 1.410 0.302 145.0 121.25 33.24 SD 1.238 0.466 137.6 118.71 30.96 n 6(1) 5(3) 6(0) 6(0) 6(0)

Sodium range 4406-7478 5531-7129 15-4254 4033-5167 907-2082 mean 5602.0 6060.8 1371.0 4625.5 1746.2 SD 1161.7 685.9 2094.6 483.9 431.5 n 6(0) 5(0) 6(0) 6(0) 6(0)

Potassium a a range 5127-6080 5270-7848 522-1075 534-3388 0-1123 mean 5592.3 6957.6 719.8 2216.3 743.7 SD 329.9 1035.7 235.9 944.1 417.3 n 6(0) 5(0) 6(0) 6(0) 6(1)

Table E3.5 cont. Summary statistics for metal concentrations (mg/kg tissue) in tissues from six turtles.

216

The dry weight of liver is on average 25% (SD = 4%) of the wet weight of the liver, and the

dry weight of kidney is on average 23% (SD = 3%) of the wet weight of kidney. So per dry

weight values reported in this study may be divided by 4 as a rough conversion to per wet

weight values for comparison with metal concentrations in other studies (Section E4). The

dry weight of femur is on average 73% of the wet weight of the femur, so dry weight values

reported in this study may be divided by 1.4 as a conversion to per wet weight values for

comparisons with other studies (Section E4). The equivalent conversion factor for claw was 1.3.

E3.5 Metals in Carapace

Eight metals (Ti / Mn, Zn / Mg, Ca, Sr, Ba / Na) were detected in all carapace samples from

wild-caught C. longicollis and E. macquarii. Almost all carapace samples showed

detectable Pb (all except one C. longicollis from Jibbons lagoon), and K (all except two C.

longicollis from Sir Joseph Banks Park). Six other metals (Al, Zr / Cr, Cu, Se, Fe) were

detected in some samples, but not others. Fourteen metals (V, Ni, Cd, Sn, Sb, As / Co, Mo /

Be / Li, Cs / La, Th, U) were not detected in any carapace sample. Results for Al, Se, and

Cu should be interpreted with caution as these metals were at concentrations near detection

limits. Carapace dry weight was on average 80% of carapace wet weight for E. macquarii

and C. longicollis from the Sludge Lagoon, and 83% for all other C. longicollis. Thus,

division of dry weight values by 1.2 can be used as a rough conversion factor to wet weight

concentrations for all carapace samples.

E3.5.1 Size

In no case is there a significant change in carapacial metal concentration (16 metals: Pb, Ti,

Zr, Al / Cr, Mn, Cu, Se, Zn, Fe / Mg, Ca, Sr, Ba / Na, K) with size (CL, Section E2.3.3) for

either C. longicollis (n = 23, CL = 126-210) from Sir Joseph Banks Park (r2 = <0.001-0.100,

p = 0.980-0.141) or E. macquarii (n = 18, CL = 133-254) from Bicentennial Park (r2 =

<0.001-0.064, p = 0.982-0.310).

E3.5.2 Sex

Sex and gravidity (Section E2.3.3) have little influence on carapacial metal concentrations

for either C. longicollis or E. macquarii. For C. longicollis at Sir Joseph Banks Park, there

are no significant differences between males and females in the carapace concentrations of

15 metals (Pb, Ti, Zr, Al / Cr, Mn, Cu, Se, Zn / Mg, Ca, Sr, Ba / Na, K). There is a

217

218

significant difference between sexes in carapace concentration of Fe (Mann-Whitney test

statistic = 30.0, p = 0.042), however, there is extensive overlap in the ranges for the two

sexes (males: 3.2-19.6 mg/kg; females: 0-12.4 mg/kg).

For E. macquarii at Botany Swamps, there are no significant differences between metal

concentrations in carapace of non-gravid and gravid female turtles in 14 metals (Pb, Ti, Zr,

Al / Cr, Mn, Cu, Se, Zn, Fe / Ca, Sr / Na, K), but Ba and Mg are on the cusp of significance

(in both cases, Mann-Whitney test statistic = 2.0, p = 0.05). With gravid and non-gravid

females pooled for all except Ba and Mg (but only non-gravid females used in the

comparison for the latter two metals), there are no significant differences between males and

females in carapace concentrations of 14 metals (Pb, Zr, Al / Cr, Mn, Cu, Se, Zn, Fe / Mg,

Ca, Ba / Na, K). Significant differences between the sexes are found for concentrations of Ti

(Mann-Whitney test statistic = 42.0, p = 0.009) and Sr (Mann-Whitney test statistic = 40.0,

p = 0.020). Gravid females have higher concentrations of Ba and Mg than males and non-

gravid females, while females usually have higher concentrations of Ti and Sr than males.

E3.5.3 Species Differences

Chelodina longicollis and E. macquarii were both caught at five sites, although only one E.

macquarii was caught at Sir Joseph Banks Park, so this site is not included in species

comparisons (Tables E3.6 – E3.9).

There are few significant differences in carapace metal concentrations between C.

longicollis and E. macquarii, and the few differences found within sites are not consistently

present across sites. At Bicentennial Park, the site with the largest samples (8 C. longicollis,

18 E. macquarii), there are significant differences between the two species in concentrations

of Ba (Mann-Whitney test statistic = 114.0, p = 0.020) and Zn (Mann-Whitney test statistic

= 23.0, p = 0.006), but not in the other 14 detected metals. At Kangaroo Creek, the two C.

longicollis have significantly different carapace concentrations to the 8 E. macquarii of Ti

(Mann-Whitney test statistic = 16.0, p = 0.037), Ca (Mann-Whitney test statistic = 16.0, p =

0.037) and Sr (Mann-Whitney test statistic = 0.0, p = 0.037), but not the other 13 metals. At

Lake Toolooma (5 C. longicollis, 3 E. macquarii), there are significant differences between

Bicentennial Park

C. longicollis (n = 8) E. macquarii (n = 18)

Metal Range Median Mean SD Range Median Mean SD

Pb 7.10-40.7 25.0 24.6 13.4 9.73-58.9 28.1 28.7 13.0

Al 0.00-10.0 0.00 2.75 4.03 0.00-1923 0.00 140 464

Ti 20.0-55.1 25.9 29.8 11.7 13.7-80.5 21.7 27.8 16.5

Zr 0.00-1.17 0.207 0.314 0.411 0.00-1.01 0.00 0.16 0.33

Cr 0.485-15.5 6.80 7.97 6.12 0.00-13.8 5.83 6.39 4.24

Mn 6.59-42.2 12.8 16.0 11.6 7.04-20.9 9.43 11.2 4.23

Cu 0.358-3.09 0.637 1.17 1.00 0.00-2.95 0.72 0.89 0.86

Se 0.00-1.60 1.21 0.852 0.716 0.00-9.44 1.34 1.90 2.28

Zn 71.0-277 91.5 110 68.5 77.0-245 112 123 35.7

Fe 4.00-47.0 11.5 17.6 16.2 6.00-82.0 12.0 18.1 18.6

Mg 2365-2855 2721 2691 149 2204-3755 2587 2656 329

Ca 217545-237712 227279 227339 7490 197402-390480 219872 228151 41735

Sr 213-691 366 406 149 308-1048 348 410 173

Ba 48.0-368 131 169 106 57.0-286 79.0 89.9 51.0

Na 4460-5828 5343 5230 514 2554-8469 4884 4839 1678

K 424-729 480 516 95.5 448-1099 537 633 190

Table E3.6 Metal concentrations (mg/kg) in the carapace of two species of turtle caught at Bicentennial Park, Rockdale. Significantly different

concentrations are in bold.

219

Botany Swamps



Pb 49.9-73.4 61.6 61.6 16.6 12.4-82.4 41.3 44.1 21.5

Al 0.00-0.00 0.00 0.00 0.00 0.00-3295 0.00 222 850

Ti 23.9-28.2 26.1 26.1 3.01 13.2-36.2 21.1 21.4 5.45

Zr 0.00-0.00 0.00 0.00 0.00 0.00-2.46 0.00 0.34 0.70

Cr 5.30-6.92 6.11 6.11 1.15 0.00-10.6 3.04 3.39 2.87

Mn 9.81-20.3 15.1 15.1 7.45 6.29-23.9 9.75 11.0 4.29

Cu 0.371-0.386 0.378 0.378 0.010 0.00-11.6 0.48 2.10 3.60

Se 1.421-1.433 1.43 1.43 0.009 0.00-8.08 1.25 1.91 2.15

Zn 105-109 107 107 2.83 97.0-217 142 147 35.4

Fe 4.00-6.00 5.00 5.00 1.41 4.00-176 14.0 28.1 43.6

Mg 2568-2866 2717 2717 211 2286-2829 2545 2528 167

Ca 244231-255554 249893 249892 8007 207586-272855 228724 231387 17769

Sr 211-832 527 527 432 257-1088 529 530 224

Ba 83.0-493 288 288 290 84.0-248 168 159 51.1

Na 5349-5385 5367 5367 25.5 3725-5769 4960 4905 590

K 380-394 387 387 9.90 307-993 497 522 193

Table E3.7 Metal concentrations (mg/kg) in the carapace of two species of turtle caught at Botany Swamps, Eastlakes. There were no significant

differences between the species.

220

Lake Toolooma



Pb 12.9-75.5 26.7 35.5 24.2 9.31-13.6 9.84 10.9 2.33

Al 0.00-2675 0.00 537 1195 0.00-0.00 0.00 0.00 0.00

Ti 22.3-74.3 23.2 42.9 27.7 19.0-22.1 21.6 20.9 1.66

Zr 0.00-0.00 0.00 0.00 0.00 0.00-0.67 0.51 0.39 0.35

Cr 2.53-14.5 5.68 6.93 4.97 0.78-4.81 3.09 2.90 2.02

Mn 6.87-47.1 12.2 18.4 16.4 6.76-11.8 11.2 9.92 2.75

Cu 0.00-0.586 0.404 0.302 0.284 0.27-0.81 0.37 0.48 0.28

Se 0.00-1.68 1.31 0.912 0.843 1.34-1.52 1.39 1.42 0.09

Zn 66.0-211 82.0 109 58.7 104-163 107 125 33.2

Fe 2.00-22.0 16.0 13.4 7.47 0.00-16.0 7.00 7.67 8.02

Mg 2511-2945 2679 2682 176 2163-2538 2532 2411 215

Ca 219966-267778 250773 251232 19421 202127-227906 204528 211520 14241

Sr 480-1025 787 739 241 603-837 776 739 121

Ba 122-598 330 324 191 263-374 355 331 59.4

Na 5181-12260 5436 7470 3130 4386-5347 4473 4735 532

K 192-452 380 325 121 353-837 400 530 267

Table E3.8 Metal concentrations (mg/kg) in the carapace of two species of turtle caught at Lake Toolooma, Heathcote National Park.

Significantly different concentrations are in bold.

221

222

Kangaroo Creek



Pb 27.3-35.9 31.6 31.6 6.07 7.95-32.8 14.7 17.8 9.67

Al 0.00-0.00 0.00 0.00 0.00 0.00-566 0.00 70.8 200

Ti 21.6-22.7 22.2 22.2 0.734 14.9-21.0 20.5 19.5 2.26

Zr 0.00-0.507 0.254 0.254 0.359 0.00-2.71 0.00 0.60 1.05

Cr 4.46-7.05 5.75 5.75 1.83 2.22-10.3 5.97 6.02 2.56

Mn 10.5-29.7 20.1 20.1 13.6 7.33-13.6 10.3 10.4 2.27

Cu 0.452-0.549 0.501 0.501 0.069 0.00-3.26 0.00 0.73 1.26

Se 1.29-1.43 1.36 1.36 0.100 0.00-2.04 1.39 1.31 0.59

Zn 95.0-172 134 134 54.4 86.0-139 108 112 18.6

Fe 5.00-33.0 19.0 19.0 19.8 4.00-54.0 22.5 24.6 15.6

Mg 2345-2865 2605 2605 368 2696-3301 2914 2927 218

Ca 233810-254097 243954 243954 14345 198939-232473 216741 217318 10599

Sr 322-673 498 498 248 850-1077 921 934 67.8

Ba 114-373 244 244 183 371-514 426 439 51.6

Na 4780-5749 5265 5265 685 4028-4862 4686 4607 266

K 431-512 472 472 57.3 431-1216 858 844 310

Table E3.9 Metal concentrations (mg/kg) in the carapace of two species of turtle caught at Kangaroo Creek, Royal National Park. Significantly

different concentrations are in bold.

223

species in Ti (Mann-Whitney test statistic = 15.0, p = 0.025), but not in the other 15 metals.

At Eastlakes (2 C. longicollis, 15 E. macquarii) there are no significant differences between

species in any of the 16 metals. The lack of consistency between sites in the metals that show

significant differences (only Ti was different at two sites), the low number of significant

differences (6 out of 68 comparisons; 8.8%), and the fact that the C. longicollis range

encompasses the E. macquarii range for both of the statistically different metals at the site

with the largest sample size (Bicentennial Park), suggests that most, or all, significant

differences between species are artefacts of multiple comparisons and small samples. Hence,

the data for the two species are pooled for comparisons between sites.

E3.5.4 Site Differences

The distribution of carapace metal concentrations for most sites, with data for the two species

pooled, is either highly skewed or platykurtic or both. Hence, comparisons among sites are

made with Kruskal-Wallis non-parametric tests.

Using the four urban, four park, and the sludge lagoon sites (Tables E3.10-E.3.18), there are

significant differences among sites for all metals except Al, Zr, Cr, and Se (Table E.3.19).

The case is the same when the sludge lagoon data are removed, except that Cr then also

shows significance. When just the four park sites are compared, there are significant

differences among sites for 7 metals (Pb, Cu, Zn / Ca, Ba / Na, K), but not the other 9 metals

(Ti, Zr, Al / Cr, Mn, Se, Fe / Mg, Sr). When just the four urban sites (U1-U4) are compared,

there are significant differences between sites for 12 metals (Pb, Ti, Zr / Cr, Mn, Zn, Fe / Mg,

Ca, Sr, Ba / K), but not the other 4 metals (Al / Cu, Se / Na).

E3.6 Metals in Blood

E3.6.1 Blood vs Environment

Metal concentrations in blood were available for C. longicollis from U3 (Sir Joseph Banks

Park) and E. macquarii from U4 (Bicentennial Park) (Section E2.3.3). For blood, dry weight

was on average 13% of wet weight (SD = 1%), so division by 7.7 will approximately convert

dry weight to wet weight concentrations for comparison with other studies. Eleven metals

(Pb, Ti / Cu, Se, Zn, Fe / Mg, Ca, Sr / Na, K) were detected in turtle blood. There are no

significant differences between blood metal concentrations at the two sites, except for Fe (U3

1164-1840 mg/kg, mean 1659.0, SD 173.5, n = 15; t = 2.747, df = 27, p = 0.011; U4 1661-

U1 Centennial Park (n = 3)

Metal Range Median Mean SD

Pb 12.7-110 81.9 68.2 50.1

Al 0.00-0.00 0.00 0.00 0.00

Ti 24.0-28.2 27.3 26.5 2.18

Zr 0.96-2.00 1.17 1.37 0.55

Cr 0.39-1.09 0.89 0.79 0.36

Mn 7.71-11.5 10.9 10.0 2.03

Cu 0.92-1.48 0.96 1.12 0.31

Se 1.45-1.78 1.74 1.65 0.18

Zn 98.0-146 135 126 25.2

Fe 31.0-93.0 39.0 54.3 33.7

Mg 2464-3013 2631 2703 281

Ca 218424-238318 220905 225882 10841

Sr 264-324 306 298 30.8

Ba 115-174 137 142 29.8

Na 4943-5698 5514 5385 394

K 238-925 768 645 360

Table E3.10 Metal concentrations (mg/kg) in carapace of turtles from

Centennial Park.

U2 Botany Swamps (n = 17)


Pb 12.4-82.4 43.6 46.1 21.3

Al 0.00-3295 0.00 196 799

Ti 13.2-36.2 21.5 21.9 5.38

Zr 0.00-2.46 0.00 0.30 0.66

Cr 0.00-10.6 3.17 3.71 2.84

Mn 6.29-23.9 9.80 11.5 4.63

Cu 0.00-11.6 0.48 1.90 3.41

Se 0.00-8.08 1.27 1.85 2.01

Zn 97.0-217 131 142 35.7

Fe 4.00-176 14.0 25.4 41.5

Mg 2286-2866 2558 2551 176

Ca 207586-272855 233598 233565 17834

Sr 221-1088 529 529 236

Ba 83.0-493 168 174 96.8

Na 3725-5769 5207 4959 572

K 307-993 470 506 186


Botany Swamps.

224

U3 Sir Joseph Banks Park (n = 24)


Pb 0.36-45.6 20.2 21.1 11.8

Al 0.00-183 0.00 10.3 36.9

Ti 16.0-75.7 25.1 30.8 15.0

Zr 0.00-2.08 0.00 0.34 0.59

Cr 0.19-13.3 6.00 6.46 3.76

Mn 3.67-18.8 7.28 8.12 3.21

Cu 0.00-2.58 0.75 0.94 0.76

Se 0.00-3.13 1.27 0.97 0.81

Zn 57.0-125 70.5 73.5 15.6

Fe 0.00-20.0 3.00 5.71 6.01

Mg 2483-3157 2915 2914 145

Ca 209715-258418 242091 240200 12495

Sr 180-1183 466 478 176

Ba 21.0-131 54.0 62.1 27.8

Na 2627-6775 5041 5009 748

K 0.00-725 361 370 173


Sir Joseph Banks Park.

U4 Bicentennial Park (n = 26)


Pb 7.10-58.9 28.1 27.5 13.0

Al 0.00-1923 0.00 97.7 388

Ti 13.7-80.5 22.2 28.4 14.9

Zr 0.00-1.17 0.00 0.21 0.36

Cr 0.00-15.47 5.83 6.88 4.82

Mn 6.59-42.19 9.66 12.6 7.40

Cu 0.00-3.09 0.72 0.97 0.89

Se 0.00-9.44 1.28 1.58 1.98

Zn 71.0-277 110 119 47.1

Fe 4.00-82.0 12.0 17.9 17.6

Mg 2204-3755 2641 2667 283

Ca 197402-390480 222417 227901 34645

Sr 213-1048 354 409 163

Ba 48.0-368 86.5 114 79.2

Na 2554-8469 5058 4959 1422

K 424-1099 526 597 173


Bicentennial Park.

225

P1 Lake Toolooma (n = 8)


Pb 9.31-75.5 18.4 26.3 22.3

Al 0.00-2675 0.00 336 945

Ti 19.0-74.3 22.5 34.5 23.9

Zr 0.00-0.67 0.00 0.15 0.28

Cr 0.78-14.53 3.95 5.42 4.43

Mn 6.76-47.1 11.5 15.2 13.2

Cu 0.00-0.81 0.39 0.37 0.28

Se 0.00-1.68 1.37 1.10 0.69

Zn 66.0-211 104 115 48.5

Fe 0.00-22.0 13.5 11.3 7.69

Mg 2163-2945 2536 2581 225

Ca 202127-267778 239154 236341 26380

Sr 480-1025 782 739 194

Ba 122-598 343 326 148

Na 4386-12260 5373 6444 2772

K 192-837 390 402 200


Lake Toolooma.

P2 Kangaroo Creek (n = 10)


Pb 7.95-35.9 20.5 20.6 10.5

Al 0.00-266 0.00 56.6 179

Ti 14.9-22.7 20.7 20.0 2.31

Zr 0.00-2.71 0.00 0.53 0.94

Cr 2.22-10.3 5.97 5.97 2.35

Mn 7.33-29.7 10.5 12.3 6.42

Cu 0.00-3.26 0.23 0.68 1.11

Se 0.00-2.04 1.39 1.32 0.52

Zn 86.0-172 108 116 26.1

Fe 4.00-54.0 22.5 23.5 15.4

Mg 2345-3301 2882 2863 265

Ca 198939-254097 220023 222645 15374

Sr 322-1077 910 847 210

Ba 114-514 408 400 112

Na 4028-5749 4723 4738 429

K 413-1216 747 769 316


Kangaroo Creek.

226

227

P4 Marley Lagoon (n = 8)


Pb 8.35-38.8 17.4 20.7 11.6

Al 0.00-925 0.00 127 323

Ti 16.1-77.0 22.4 41.3 28.9

Zr 0.00-1.16 0.00 0.21 0.42

Cr 1.55-8.82 3.70 4.68 2.66

Mn 6.21-39.8 6.86 12.9 12.0

Cu 0.00-0.69 0.35 0.31 0.28

Se 0.00-3.97 1.46 1.29 1.34

Zn 74.0-95.0 80.5 83.3 8.51

Fe 1.00-55.0 9.00 15.6 17.7

Mg 2281-3150 2673 2696 301

Ca 224419-273598 259157 255821 18477

Sr 470-998 753 743 167

Ba 50.0-458 95.0 153 139

Na 2615-21288 5534 7182 5811

K 147-560 231 266 126


Marley Lagoon.

P3 Jibbons Lagoon (n = 8)


Pb 0.00-24.7 7.59 7.71 7.70

Al 0.00-18.0 12.5 9.50 8.14

Ti 16.4-35.0 29.2 27.8 7.56

Zr 0.00-0.00 0.00 0.00 0.00

Cr 0.00-8.83 5.02 4.22 3.65

Mn 3.40-47.2 6.80 11.6 14.6

Cu 0.56-3.59 2.07 1.87 1.04

Se 0.00-3.89 0.00 0.90 1.42

Zn 81.0-105 95.0 94.5 8.14

Fe 7.00-20.0 11.5 12.0 4.34

Mg 2388-2981 2771 2738 168

Ca 189675-273877 244198 240017 23578

Sr 314-1038 560 750 258

Ba 76.0-557 217 254 166

Na 3284-6092 5744 5495 914

K 142-1532 352 464 443


Jibbons Lagoon.

228

Sludge Lagoon (n = 11)


Pb 1.79-43.5 11.7 16.0 13.8

Al 0.00-431 0.00 43.6 129

Ti 14.0-77.0 21.5 33.4 22.9

Zr 0.00-1.86 0.00 0.40 0.60

Cr 0.00-7.61 4.41 4.38 2.45

Mn 2.78-13.8 4.60 5.41 3.09

Cu 0.00-3.52 0.53 0.69 0.98

Se 0.00-6.00 1.32 1.43 1.65

Zn 59.0-196 92.0 95.3 35.5

Fe 11.0-118 38.0 52.8 34.5

Mg 2193-3756 2858 2867 380

Ca 201585-266454 226513 224730 19782

Sr 239-492 318 323 68.0

Ba 15.0-175 24.0 61.0 62.5

Na 2401-5241 4665 4323 1028

K 326-941 634 596 176

Table E3.18 Metal concentrations (mg/kg) in carapace of turtles from the sludge lagoon.

229

Four park sites

(df = 3)

Four urban sites

(df = 3)

Four urban + four park

sites (df = 7)

All nine sites

(df = 8)

Metal K-W statistic p K-W statistic p K-W statistic p K-W statistic p

Pb 11.438 0.010 15.566 0.001 32.111 <0.001 35.768 <0.001

Al 3.936 0.268 3.394 0.335 9.352 0.228 9.033 0.339

Ti 6.439 0.092 7.943 0.047 16.065 0.025 15.744 0.046

Zr 3.786 0.286 9.365 0.025 14.094 0.050 14.802 0.063

Cr 1.575 0.665 12.171 0.007 14.510 0.043 15.167 0.056

Mn 6.623 0.085 13.972 0.003 20.422 0.005 35.211 <0.001

Cu 13.471 0.004 2.058 0.560 21.788 0.003 23.347 0.003

Se 0.934 0.817 4.312 0.230 6.243 0.512 6.376 0.605

Fe 4.170 0.244 22.166 <0.001 27.816 <0.001 42.436 <0.001

Zn 12.877 0.005 40.452 <0.001 57.803 <0.001 61.683 <0.001

Mg 6.146 0.105 30.324 <0.001 36.779 <0.001 38.574 <0.001

Ca 9.237 0.026 17.390 0.001 28.287 <0.001 30.911 <0.001

Sr 2.930 0.403 12.983 0.005 44.167 <0.001 54.630 <0.001

Ba 10.582 0.014 32.493 <0.001 60.013 <0.001 65.051 <0.001

Na 8.316 0.040 1.421 0.701 15.742 0.028 20.848 0.008

K 14.424 0.002 18.207 <0.001 38.215 <0.001 41.925 <0.001

Table E3.19 Statistical comparisons of metal concentrations in carapace among sites. Significant differences are in bold.

2137 mg/kg, mean 1821.4, SD = 141.9, n = 14). Fe was not detected in the water at either of

the sites, but sediment concentrations of Fe were higher (8301 mg/kg) at U4 than at U3 (2666

mg/kg).

E3.6.2 Blood vs Carapace

As fresh whole blood samples were not available from the six dead turtles used for tissue

metal comparisons, metal concentrations in blood with carapace were compared in C.

longicollis from Sir Joseph Banks Park. The sample size at this site (5 males, 7 females) is

sufficient for testing for sexual differences in metal concentrations. Because the distributions

of concentrations are zero-truncated for some metals, all comparisons between sexes are

made with Mann-Whitney U-tests. There are no significant differences in metal (Pb, Ti / Cu,

Se, Zn, Fe / Mg, Ca, Sr / Na, K) concentrations between male and female C. longicollis from

Sir Joseph Banks Park in either blood or carapace (p = 0.223-0.953), except for Fe in

carapace (males: 3.2-19.6, mean = 12.7, SD = 6.22, n = 5; females: 0-12.9, mean = 4.7, SD =

5.11, n = 7; U = 30, p = 0.042).

Because of the lack of significant sexual differences, the sexes are pooled (Table E3.20) for

exploration of correlation between blood and carapace metal concentrations, except for Fe,

where the comparison is restricted to females. There are no significant correlations between

blood and carapace metal concentration for any of the 11 metals (Pb, Ti / Se, Cu, Zn, Fe /

Mg, Ca, Sr / Na, K) (Pearson’s p = -0.458-0.551; p = 0.089-0.998).

Correlation between metal concentrations in blood and carapace is also tested for E.

macquarii at Bicentennial Park, where the largest sample was taken. Blood data are available

for one male and seven female E. macquarii, so analysis of sexual differences in metal

concentrations is not made. However, the single male value falls within either the range of

female values or the 95% confidence interval for the mean female value, or both, for most

metals and tissues. There are seven exceptions (Table E3.21). For Pb, Ti, Mg, Ca, Sr, Na and

K, correlations between metal concentrations in blood and carapace are restricted to the

female E. macquarii sample (n = 7), while for Se, Cu, Zn and Fe, the correlations used

pooled sexes (n = 8) (Table E3.22).

There are no significant correlations between metal concentrations in blood and carapace for

Ti, Se, Cu, Fe Mg, Ca, Sr, Na or K (Pearson’s p = -0.400 - 0.635; p = 0.125-0.745). There is

230

Carapace Blood

Metal n Range Mean SD Range Mean SD

Pb 13 1.6-45.6 26.1 12.67 0-0.64 0.19 0.226

Ti 13 16.0-74.2 28.1 14.62 0-35.2 7.38 10.05

Cu 13 0-2.2 0.8 0.66 3.95-5.71 4.70 0.62

Se 13 0-3.1 1.2 0.84 2.52-6.41 3.85 1.21

Zn 13 56.7-125.3 77.1 19.46 26.9-41.0 34.09 4.43

Fe 7 0-12.4 4.7 5.11 1445-1840 1668.9 136.2

Mg 13 2483-3043 2863.3 143.12 288-473 352.8 42.25

Ca 13 209715-258418 241471.4 14077.85 611-979 746.2 110.99

Sr 13 180-638 444.2 104.43 0.56-8.79 2.29 2.04

Na 13 2627-6775 5027.5 926.4 13701-19373 16850.5 1771.85

K 13 0-725 393.8 180.3 4909-6787 6119.1 516.17

Table E3.20 Data used for comparison of metal concentrations (mg/kg) in carapace and blood of C. longicollis from Sir Joseph Banks Park (U3). Sexes pooled for all but Fe, where data are for females only.

Female Values Female Con. Lims

Metal Tissue

Male

Value Min. Max. Lower Upper

Pb blood 0.420 [0.420] - 0.571 -

Mg blood 356 603 - 593 -

Ca blood 514 603 - 593 -

Sr blood 4.92 - 2.68 - 2.41

K blood 6815 - 6598 - 6372

Ti carapace 13.7 14.1 - [13.5] -

Na carapace 2554 2564 - 2819 -

Table E3.21 Sexual differences in blood and carapace metal concentrations (mg/kg) in E. macquarii from Bicentennial Park (U4), showing values for the single male that lie outside the female range and/or female confidence limits (Con. Lims). Bracketed female values are inclusive of the male value.

231

a significant correlation between blood and carapace Pb concentrations (Pearson’s p = 0.899,

p = 0.006), and correlation for Zn is on the cusp of significance (Pearson’s p = 0.707, p =

0.050).

Carapace Blood

Metal n Range Mean SD Range Mean SD

Pb 7 9.7-55.2 27.9 14.77 0.42-1.46 0.88 0.34

Ti 7 14.1-33.6 20.7 7.82 0-15.1 4.87 5.28

Cu 8 0.34-2.72 1.01 0.75 2.99-4.03 3.53 0.42

Se 8 0-3.75 2.04 1.58 1.58-2.15 1.88 0.20

Zn 8 108.3-162.1 118.2 18.14 24.1-35.4 30.03 3.54

Fe 8 6.13-43.59 13.5 12.49 1661-1996 1787.5 123.74

Mg 7 2204-2719 2507.7 171.30 295-370 329.6 25.92

Ca 7 209631-

230298

220249.6 6570.38 603-996 721.7 138.28

Sr 7 308-609 396.4 102.6 1.10-2.68 1.89 0.56

Na 7 2564-5325 4008.4 1285.56 13273-17649 15694.1 1451.96

K 7 448-1099 654.9 238.47 5660-6598 6034.6 364.46

Table E3.22 Data used for comparison of metal concentrations (mg/kg) in carapace and

blood for E. macquarii. Where n = 8 sexes are pooled, where n = 7 only females are used.

E3.7 Bone Structure

Backscattered scanning electron microscopy (SEM) of marginal carapace bone mounts

showed large canals and surrounding areas of remobilisation (Figure E3.2a), and a sharp

delineation between the keratin scute and the underlying bone (Figure E3.2b). X-ray emission

SEM showed Ca, P and O to be dominant in bone, although sulfur was also detected in the

bone seam (Figure E3.2c), and in the scutes. Scute keratin may have become lodged in the

seam or be there naturally. Microscopy showed no sign of lamination of carapacial bone. The

bone was pitted and uneven and unsuitable for multiple surface analysis for metals using

Secondary Ion Mass Spectrometry (Rob Russell pers. comm.).

232

The edge of the bone of the marginal scutes was hard and compact, but became

predominantly spongy within 6 mm of the edge (Figure E3.3a). Femur, in contrast, had very

little compact bone (Figure E3.3b). The marginal bone at the edge of the E. macquarii

carapace was not examined but is much thinner and more maleable than the harder and

thicker edge of the C. longicollis carapace.

233

CHAPTER E4: METAL BIOACCUMULATION DISCUSSION

E4.1 Turtle Captures

Chelodina longicollis is widespread in southern Sydney, having been captured at all

urban (n = 4) and park (n = 4) sites. Emydura macquarii is also present at both urban

(n = 3) and park (n = 2) sites, whereas Elseya latisternum is only found at urban (n =

2) sites. As found in the Sydney Survey (Sections C3.1, C3.2), E. macquarii was not

found at any sites where C. longicollis was not present (Sections D3.1.2, D3.1.5),

again indicating that, as in natural populations (Chessman 1978, Cann 1998), they do

not displace the latter when introduced into urban environments.

Only small remnants of bushland still exist in Sydney’s highly urbanised and

industrialised southeastern suburbs (location of urban sites); for example, only 0.6%

of the original bushland area in the suburb of Rockdale (Urban 4) remains (Benson &

Howell 1990). Similarly, freshwaters and swamps are greatly reduced from their

original extent, with land reclaimed for housing, industry and airport construction

(Benson & Howell 1990). Nevertheless, turtles maintain viable populations in these

reduced and isolated habitats. Although the size (and thus probably the age) of turtles

varies among sites, their body condition does not, suggesting that there are adequate

resources for normal growth at all sites. Normal weight, however, does not indicate an

absence of metal effects, as the weight of aquatic animals may be the same at metal-

polluted sites compared to control sites, even when tissue metal concentrations are

elevated (Dethloff et al. 2001).

The sparsity of urban freshwater lagoons means that large accessible lagoons are the

likeliest places for illegal release of unwanted pet turtles. Emydura macquarii is

introduced at Botany Swamps (U2), as in the late 1940s only C. longicollis was

present (Mackay 1949). The appeal of this site for release of turtles is confirmed by

observations of other translocated species nesting there (C. expansa, Cann 1998).

Chelodina longicollis is the only species considered to occur naturally at Centennial

Park (U1) (Stephenson 1986), even though E. macquarii is also present (Sections

C3.1, C3.2). Although numbers are low (n = 3), captures of only C. longicollis in

Model Yacht Pond (U1) are consistent with previous observations that this is the only

species presence in the smaller, shallower ponds of Centennial Park (Stephenson

235

1986). Emydura macquarii is also likely to have been introduced into the Royal

National Park, as prior to the construction of Audley Weir, Kangaroo Creek (P2) and

the connected Hacking River would have been much smaller and similar to other river

systems to the south of Sydney where only C. longicollis is present (Griffiths 1997).

E4.2 Environmental Parameters & Urbanisation

E4.2.1 Physicochemical Parameters

pH

Lowland rivers (Kangaroo Creek) and freshwater lakes (all other sites) are stressed if

their pH is outside the range of 6.5-8.0 (ANZECC & ARMCANZ 2000a), which is

the case at Sir Joseph Banks Park (U3) and Bicentennial Park (U4) (high), and Lake

Toolooma (P1) and Kangaroo Creek (P2) (low). The draining of wetlands along much

of the NSW coastline, especially for urban development, has caused oxidation of acid

sulfate soils with release of sulfuric acid into surface waters (EPA 1996). However,

the highly disturbed urban sites all had higher pH values than all the park sites, and it

is not known if urban (e.g. industrial wastes) or natural (e.g. soil type) phenomena are

responsible. The affect of pH requires further study as changing pH alters metal

availability, but not in a consistent fashion over metals (Mason 1996), making it

difficult to separate the direct effect of decreasing pH from changing metal

concentrations (Green 2001).

Acidic waters cause the destruction of native macrophytes and subsequent invasion by

acid-tolerant species such as Eleocharis spp. (reviewed in EPA 1996), which was

found at all park, but no urban sites. Changes to diet (and thus gastrointestinal

pollution exposure) or behaviour of Australian freshwater turtles caused by plant

community alteration is not known. Turtles in this study were found living in waters

of pH 5.7-9.0, whereas Australian turtles have previously been found in waters from

pH 5.5 to 8.2 (Cann 1998). Hence, pH tolerance is very broad, although the extremes

of tolerance, and the effect of longterm extreme pH exposures, remain unknown.

236

Salinity

Salinity in freshwaters has a wide natural range (20-2200 μS/cm), but is usually low

in lakes (20-30 μS/cm), and mid-range in NSW coastal rivers (200-300 μS/cm)

(ANZECC & ARMCANZ 2000a). All the lagoons have higher salinity than expected

for freshwater lakes (Table E3.2), possibly due to their coastal nature. Only Marley

Lagoon is above the range for natural freshwaters, a result of intermittent channel

connection with the sea. The influence of water salinity on metal accumulation in

turtles requires further examination, as liver and kidney metal concentrations are

similar in turtles from uncontaminated and metal-contaminated freshwaters, but

higher in turtles from metal-contaminated brackish water (Chelydra serpentina,

Albers et al. 1986). In the current study, the high salinity at Marley Lagoon may put

turtles at this site at greater risk of high metal exposure (Section E4.5.1 Sites &

Urbanisation).

Oxygen

As aquatic oxygen concentrations decrease, respiratory rates in fish are elevated,

leading to increased uptake of dissolved pollutants (Mason 1996). Metals are taken up

by accessory respiratory structures in freshwater turtles (Jeffree 1991, reviewed in

Jeffree & Jones 1992) which significantly influences toxin absorption (Palmer 2000),

although this needs quantification. Switching from aquatic to atmospheric respiration

may also be protective. As many fish species are adversely impacted when dissolved

oxygen concentrations fall below 5 mg/L (ANZECC & ARMCANZ 2000a) fish

abundance and hence turtle diet could be affected, however, all eight sites have

oxygen concentrations higher than this.

Turbidity

Carp (Cyprinus carpio), which were detected at all urban but no national park sites,

increase turbidity in freshwater systems (Bales 1994). Clay soils, which are common

in the Sydney area, do not aggregate well and so also increase turbidity (EPA 1996).

Due to their strong affinity for clay particles, metals may have a lower availability in

clay-based aquatic systems (Pacyna 1996), but ingestion of suspended particles during

drinking or feeding may lead to the release of bound metals in the low pH

environment of the stomach (Luoma 1983), as has been suggested for other aquatic

237

reptiles (Crocodylus porosus, Twining et al. 1999). In this way urban turtles could be

exposed to elevated metals by increased turbidity resulting from introduction of carp,

especially around clay sediements.

Depth Profiling

A one off water quality depth profile (surface to base of 4.5 m) was conducted at

Bicentennial Park (U4, Rockdale City Council pers. comm.), and points to the

complexity of physicochemical parameters at field sites. Electrical conductivity shows

salinity stratification with the less dense, less saline water on top, a marked increase in

salinity over the surface one metre, and highly saline water below this (three times

higher than at the surface). The marked salinity stratification allowed temperature to

increase with depth (9°C to 15°C), and the water was anoxic below one metre. Thus,

even multiple analysis of surface water may not adequately reflect the conditions

present in most of the water column. This emphasizes the need for field studies to

determine the details of aquatic habitat use, in combination with laboratory studies

examining metal uptake and effects under defined conditions that reflect field

exposures.

E4.2.2 Environmental Metals

Hypothesis 5 (Section E1.2), that urban waterbodies do not differ in metal

concentrations to non-urban waterbodies, is not rejected. While urban sites as a whole

do not have greater or lesser concentrations of metals in water or sediment than non-

urban sites, there is much individual variation in metal concentrations across sites.

Some urban (and some non-urban) sites have very high concentrations of some

metals.

Of the 10 metals not detected in any environmental or turtle sample (Section E3.2.2),

Hg is of greatest environmental concern (e.g. Zillioux et al. 1993). Hg is volatile and

an alternate analytical technique such as cold vapour atomic absorption spectroscopy

is required for adequate detection capability (David Hill pers. comm.). As Hg has

previously been detected in turtle muscle (Helwig & Hora 1983), kidney (Meyers-

Schöne et al. 1993), and scutes (Presti et al. 1999), and has adverse effects on the

238

immune (Bagenstose et al. 1999), and reproductive (Lundholm 1997) systems, it

should be included in future studies examining metal impacts on turtles.

Of the 12 metals detected in the environment but not in any turtle samples (Section

E3.2.2), Co is the only essential trace element. Its lack of detection is unusual as it is

an essential component of vitamin B12 (Elinder & Friberg 1979), and is usually at

highest concentrations in liver, kidney, and bone (Underwood 1977), all of which are

analysed in this study. Of the non-essential elements, Cd and As are the ones of most

ecotoxic concern (Goyer & Cherian 1995). The detected sediment concentrations of

Cd at U4 and P3 (0.6 and 0.9 mg/kg) are slightly above the normal range for

freshwater sediments (0.05-0.5 mg/kg) (Jensen & Bro-Rasmussen 1992), but Cd does

not biomagnify (Mathis & Cummings 1973, Ward et al. 1986) and most uptake is

directly from water (Beijer & Jernelöv 1979b, Jensen & Bro-Rasmussen 1992), in

which it was not detected. Manure and fertilisers often contain Cd which can then

contaminate waters receiving run-off (Jensen & Bro-Rasmussen 1992), but Cd was

not detected in water or sediment at U2 which is nestled between golf courses which

receive regular fertilisation. The absence of detectable Cd in target organs in

Australian freshwater chelids is of interest, as this metal is considered ubiquitous in

some other turtle species (Ch. mydas, Gordon et al. 1998), and has been detected

previously in freshwater turtles (Trionyx spinifer, Robinson & Wells 1975; Ch.

serpentina, Helwig & Hora 1983, Albers et al. 1986; Chrysemys picta, Rie & Callard

1997). Its absence indicates a negligible body burden, as liver and kidney are major

sites of Cd deposition in turtles (C. picta, Rie et al. 2001). For As, concentrations are

highest in marine organisms, so sea turtles show the highest concentrations amongst

the reptiles (Linder & Grillitsch 2000), but even then As may only be detected in a

small percent of the population (Aguirre et al. 1994), and may not be as problematic

within the Chelonia as it is in other taxa.

Metals in Water

Metals in freshwaters occur in many forms, and it is only the bioavailable proportion

that is of toxicological relevance (Beijer & Jernelöv 1979b, Pacyna 1996, Linder &

Grillitsch 2000), so trigger values (ANZECC & ARMCANZ 2000b, Table E3.3) may

be unnecessarily sensitive if the metal is largely unavailable for uptake by aquatic

organisms (Markich et al. 2001). Metals, especially those that are metabolic analogs

239

of Ca, are less toxic in hard waters, due to competition with Ca for uptake sites

(ANZECC 1992, Jeffree et al. 1995). Also, a deficit of dietary Ca can be of more

concern to health and reproduction than degree of metal contamination, due to its

protective role, especially at acidified sites (reviewed in Eeva & Lehikoinen 1996). So

the urban sites with very hard water (U1 & U2, Section E3.3.2) may be afforded some

protection from their dangerously elevated concentrations of Pb, Cu, Zn and Fe (Table

E3.3). Pb toxicity is also reduced by the presence of Zn (Chang 1996a), but with both

metals elevated above trigger values at U1 and U2 it is not known whether any

protective capacity of Zn is outweighed by possible direct toxic effects. Elevated

aquatic metal concentrations may be of more concern in the colder months if turtles

are active but not feeding, as starved animals accumulate metals more rapidly from

solution than do feeding animals (reviewed in Luoma 1983). This phenomenon still

needs testing in turtles.

Ba is significantly higher at urban sites (max. of 47 μg/L) compared to park sites,

although all Ba values are well within the range for natural freshwaters (reviewed in

Reeves 1979). Although Ba ions are readily absorbed from the gastrointestinal tract,

the metal has a low degree of toxicity, and concentrations of 5 mg/L in drinking water

are not toxic to vertebrates (Underwood 1977, Reeves 1979, Ware 1988a), so the

ecotoxic relevance of elevated Ba is minimal. Urban sites have the majority of higher

aquatic metal concentrations, yet there is a lack of significant difference between

urban and park sites for all metals apart from Ba. This indicates that urbanisation does

generally lead to increased aquatic metal concentrations, but that the divide between

urban and park sites may be complicated by anthropogenic impacts other than

urbanisation such as dumping, or by an underlying natural variation in conditions (e.g.

soil type). The isolation of some urban sites (U3, U4) from inputs affecting other sites

(U1, U2) is also demonstrated by differences in aquatic metal concentrations.

Pb, Mn, Cu and Fe were all detected in lagoon water from Centennial Park (U1) and

Botany Swamps (U2), but at no other sites and, except for Mn, all concentrations

exceeded recommended trigger values (Table E3.3). U1 and U2 have large entry

drains which carry wet weather runoff from the surrounding urban surfaces (as does

U4, but not U3). Stormwater runoff entering the Botany Swamps is fairly

240

characteristic for Sydney (Young 2000). Open space occupies about 48% of the

Botany wetlands catchment with the remaining area divided between industrial

(15%), residential and commercial properties, with a third of the catchment covered

with impervious surfaces (Young 2000). Urban runoff entering and flowing through

the Botany Swamps over three years exceeded recommended levels for Pb in urban

waters (20 μg/L) in 22-71% of cases (Sim 1993). Pb (in the environment and in living

tissue) is one of the metals most associated with urbanisation (Underwood 1977,

Tsuchiya 1979, Eisenmann & Miller 1996, EPA 1996), and, of ecotoxicologically

significant metals, it is the one with the highest anthropogenic:natural output ratio

(reviewed in Prosi 1989). The only correlation in metal concentrations between

environmental and carapacial samples is for Pb in water and carapace, and Pb is one

of only two metals (the other being Se, not detected in water) whose concentration in

carapace varies with water hardness (Section E3.3.4). It is unusual that the high

aquatic Ca concentrations are associated with increased Pb in carapace, although this

may be an artefact of high Ca and Pb concentrations in water at the same two sites,

and generally low concentrations of both at the remaining sites. Low dietary input of

vitamin D leads to higher bone Pb levels (Hu 1998), and this may also be a

contributing factor.

At two urban sites metals have previously been detected that were not detected in this

study, or were detected at higher concentrations than in this study (Botany Bay City

Council 1999; Botany Bay City Council pers. comm.). Water samples taken (in

1998/1999) from Sir Joseph Banks Park (U3) and surrounding sites showed

concentrations (μg/L) of up to 72 Cd, 84 Cr, 50 Cu, 69 Pb, and 8000 Zn. The Botany

Bay catchment (especially the northern end where U3 is situated) produces 55% of all

Sydney’s liquid wastes (Butlin 1976). Water testing (in 1996-1999) at two stormwater

drains entering the Botany Swamps and three surrounding drains showed metal

concentrations (μg/L) up to 6 Hg, 100 Cd, 100 Cu, and 1200 Zn. Significant

quantities of Cr, Pb, Fe and Cd also enter the Botany Swamps in leachate from

landfill (Sim 1993). The snapshot view obtained by one-off water sampling may

identify the presence of problematic metals but does not adequately assess metal

presence either spacially or temporally.

241

If mixing in lagoons is poor, aquatic organisms will be exposed to localised areas of

high metal concentration, such as at drain entry points, making correlation of metal

concentrations in water and animal tissues problematic. Turtles take up and store

water in the cloaca and bladder (Alderton 1993), so contact with highly contaminated

water picked up in one area may continue to occur over an extended period of time.

Uptake of metals from water by plants evens out short term fluctuations in dissolved

metal concentrations (Jones et al. 1985), and isolates the bioavailable fraction, so

analysis of plant samples (especially bryophytes, reviewed in Mason 1996) may be a

preferable alternative to water sampling for comparison to animal tissue metal

concentrations in the future.

Metals in Sediment

Sediment properties such as organic content, clay content, grain size, porosity, and

cation exchange capacity all alter the metal-sequestering capacity of sediments (Ward

et al. 1986, Prosi 1989, ANZECC 1992), and consequently the bioavailability of

metals. The extent to which sediment-associated contaminants can move through

aquatic food chains and potentially affect organisms at higher trophic levels must be

understood in order to implement appropriate environmental management (Suedel et

al. 1994), and this should be established for urban Sydney sites with elevated levels of

potentially toxic metals. All metals found in turtles are also found in lagoon

sediments, except for Se. This metalloid is easily absorbed gastrointestinally from

both food and water (Glover et al. 1979). As it is not detectable in lagoon water,

turtles are likely to be obtaining Se from dietary items, or to have stored Se during

periods of higher environmental availability.

Metals are concentrated in sediments and associated bacteria, and ingestion of

sediment is a major route of exposure to some environmental metals (reviewed in

Fisher 1995, reviewed in Kirby et al. 2001), and may lead to the high Pb

concentrations found in tadpoles inhabiting highway drainages (reviewed in Rader

1998). This exposure route is also applicable for freshwater turtles, with sediments

being ingested directly during feeding (Moll & Legler 1971, Spencer et al. 1998).

Emydura macquarii ingests sand and leaf litter in the Murray Valley (Spencer et al.

1998), with 6% of the stomach content volume attributable to inorganic sediment

(Chessman 1986). As with suspended particles, metals may be released from sediment

242

particles in the acid environment of the stomach. Benthic macroinvertebrates may

accumulate high quantities of metals in polluted environments (reviewed in Nagle et

al. 2001), sometimes having higher metal concentrations than any of the other biota or

the sediment in the ecosystem (Enk & Mathis 1977), so their consumption as dietary

items (Section A1.3.3) may also add to the relevance of sediment metal

concentrations. Benthic macroinvertebrates are severely degraded in metropolitan

areas in direct relation to urban density (Walsh et al. 2001), and it is likely that only

pollution-tolerant species remain. With dietary metal concentrations often correlated

to tissue metal concentrations (Reid & Hacker 1982), variations in diet among sites

should be assessed. For many Australian freshwater turtles, little is known of their

diet even in undisturbed areas, and nothing is known of urban impacts on diet and the

interaction of this with pollutant exposure.

Lead

The natural background range of Pb in sediments is 20-34 mg/kg (reviewed in Prosi

1989). All sites are within or below this range except for U1 (219 mg/kg), U2 (156

mg/kg), and P3 (737 mg/kg) (Table E3.4). Water hyacinth (Eichhornia crassipes)

from another pond near U1 (Kensington Pond, Centennial Park, Section C2.1.3), had

root Pb concentrations (145-1110 mg/kg) that intersected the range of sediment Pb

concentrations (590-1870 mg/kg) (Vesk & Allaway 1997), showing that the

sequestering of toxic metals in the biota may be considerable. Pb concentration in

sediments and roots decreased exponentially with distance from a stormwater inlet,

emphasising the need for multiple spatial assessment of metal concentrations (Vesk &

Allaway 1997). The persistence of Pb contamination in sediments for decades

(Zarcinas & Rogers 2002), and the possibility of either chronic or acute metal input,

makes it difficult to determine the nature and duration of the Pb contamination at the

Sydney sites.

Contamination of Urban and Park Site

Urban sites had the majority of highest sediment concentrations for 15 metals, and the

park sites for none, yet the urban and park sites did not differ significantly in the

concentration of any of the 27 sediment metals for which complete data are available.

This is due to the high variation in metal concentration over all sites, the overlap in

values between urban and park sites, and the presence of extremely high

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concentrations of metals at some sites (e.g. Pb at P3, Table E3.4). Pb and Zn occur at

significantly higher concentrations in soils in urban areas (median Pb 51 mg/kg, Zn 58

mg/kg) compared to neighbouring rural areas (median Pb 18 mg/kg, Zn 18 mg/kg)

(Adelaide, Tiller 1992). However, the natural lithology of a catchment is also a major

factor in local sediment composition (Geary 1981), so, to better gauge the degree of

anthropogenic metal contamination of sediments, natural variation in sediment metals

could be determined by analysis of a depth sediment sample (e.g. at 300 mm), in

addition to surface sediment analysis (Kirby et al. 2001).

Many anthropogenic factors are likely to contribute to spatial variations in metal

concentrations. For instance, atmospheric pollution causes widespread contamination

of waterbodies, especially non-flowing ones (Prosi 1989). The Royal (P2, P3, P4) and

Heathcote (P1) National Parks border Sydney to their north and the Illawarra district

to their south. The Illawarra is a narrow coastal plain that has supported heavy

industry (including steelworks, copper smelting, quarries, cokeworks, and collieries)

and associated urban development for over 100 years, and greater quantities of

industrial emissions emanate from this region than from Sydney (EPA 1996). Much

of the industry is located in Wollongong, just to the south of the parks. The Illawarra

may be affecting environmental metal concentrations at park sites in two ways.

Firstly, rubbish dumping is one of the major management problems for the Royal

National Park (Benson & Howell 1990). Historical and current disposal of industrial

and hazardous waste originating in the Illawarra is problematic (EPA 1996), so

illegally dumped waste may be the cause of elevated metal levels at some park sites

(e.g. the Pb at P3).

Secondly, there are historical and current problems with the management of air

pollution in the Illawarra, especially that originating from industrial emissions (EPA

1996). Up to 50% of metals (including Pb) may enter freshwaters via atmospheric

deposition (with, in some instances, over 30% estimated to originate from over 200

km away), and, on a global scale, anthropogenic emissions have greatly disturbed the

geochemical cycles of metals such as Pb, Cd, Se and Zn (reviewed in Pacyna 1996).

Hence atmospheric deposition of pollutants over small and large distances is an

important pathway for ecosystem contamination. In addition, the national parks are

exposed to atmospheric pollutants from Sydney that are blown out to sea at night and

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then back the following day to the south of the city (Benson & Howell 1990).

Consequently, atmospheric deposition may account for some rise in environmental

metal concentrations at park sites. However, the Botany Bay catchment (especially

the northern end where U3 is situated) has at times been responsible for producing

60% of all Sydney’s atmospheric wastes (Butlin 1976), and in the mid-1970s (10

years prior to the banning of new cars requiring leaded petrol) it was estimated that

vehicles in the Botany Bay area emitted two tonnes of Pb every day (Butlin 1977).

With this long history of intense atmospheric pollution, it is difficult to explain why

U3 has lower environmental Pb concentrations than any of the park sites, so there

must be other complicating factors.

In addition to industrial processes and fuel combustion, major contributors to

atmospheric particles are vegetation (e.g. pollen), soil dust (including bushfire ash),

and sea spray (Nriagu 1979, EPA 1996), with the latter two amongst the largest

natural sources of atmospheric metal deposition (reviewed in Linder & Grillitsch

2000). In Sydney, fine atmospheric particles include 7% sea spray and 6% soil

(Cohen et al. 1994, as cited by EPA 1996). Bushfires are significant contributors to

metals in the atmosphere (Beijer & Jernelöv 1979a), and the southeast coast of

Australia (encompassing Sydney), in contrast to the rest of the country, has a bushfire

frequency of < 5 years (Young 2000). Sydney’s national parks are often the sites of

these fires which, together with their exposure to sea spray, may explain some of the

higher metal levels in park lagoons. Turtles may also be exposed to atmospheric

metals directly by inhalation (Glover et al. 1979, Norseth 1979, Pacyna 1996), but the

extent of this route of intake is unknown.

E4.3 Biomonitoring

Non-lethal estimation of metals in internal organs

Most studies of metals in aquatic systems have focussed on fish or molluscs, with

little information on accumulation or effects of metals in aquatic reptiles. Studies

comparing metal concentrations over various organs are particularly scarce, yet organ

correlations need to be well understood in order to use non-lethally samplable tissues

for assessment of internal organ metals (Linder & Grillitsch 2000). With only limited

correlations between metal concentrations among the five tissues sampled from C.

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longicollis, and no correlations for the more ecotoxicologically important metals (Pb,

Cu, Cr), the hypothesis that carapace analysis is useful in reflecting metal

concentrations in internal target organs such as the liver and kidney is largely

discounted (Hypothesis 1 not rejected, Section E1.2), although the small sample size

means only very high correlations would be detected. The result contrasts with that

found for snapping turtles (Chelydra serpentina) where Pb concentration in carapace

correlated with liver Pb concentrations, increasing dramatically with increasing

environmental Pb contamination (Overmann & Krajicek 1995).

The correlations that do exist for C. longicollis are most often among the three bone-

containing tissues (carapace, femur, claw; for Mn, Zn, Sr, Ba), so, not surprisingly,

carapace sampling appears most suited to the detection of primarily bone-depositing

metals. Bone is a good indicator of environmental (water and sediment) Zn

concentrations as well (fish, Miller et al. 1992). Their predominance in bone may

explain why carapacial Mn and Zn correlated with whole body burdens in sea turtles

(Sakai et al. 2000b), as bone loads may swamp those in other tissues. Of the tissues

analysed from dead turtles, carapace is the only one suitable for sampling from free-

ranging animals, although claw can also be sampled non-lethally. The potential for

external tissues to reflect internal soft tissue concentrations of metals in this study is

limited to claw for the detection of Fe in liver, and carapace for the detection of Sr in

liver, metals which are not those of most toxicological interest.

All alkaline earth metals tend to accumulate in bone (Reeves 1979), with over 90% of

both Sr and Ba occurring in the skeleton (Underwood 1977). Neither metal is

considered essential, or toxic – except for inhalation of Ba in industrial settings and

exposure to 90Sr created during nuclear fission (Underwood 1977), with the latter

known to accumulate in the turtle carapace (Jackson et al. 1974). Despite expectations

to the contrary (Section E3.4, Jeffree et al. 1993), there are no correlations of Ca

concentration with either Sr or Ba concentration, except between Sr and Ca in liver

and between Ba and Ca in kidney. The reason for this is not known.

The lack of correlation across tissues for Pb contrasts with that found previously for

turtles, where Pb concentrations correlate between liver and carapace, and between

femur and carapace (Chelydra serpentina; Overmann & Krajicek 1995), so the

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applicability of carapacial biomonitoring may vary among species and habitats. As Pb

has a biological halftime of a few weeks in soft tissues and of 10-20 years in bone

(humans, Clarkson 1979, Tsuchiya 1979), correlations of Pb between tissues will also

vary with the ratio of current to historical Pb exposure. Scutes or scales were not

analysed in this study, but Pb has higher concentrations in urban compared to rural

areas for both cobra and wall lizard scales (Kaur 1988), with skin shedding in snakes

also suggested as a possible method of Pb excretion (due to its higher concentration in

skin than the rest of the body, Burger 1992). Freshwater turtle scute analysis could be

examined as a possible Pb monitoring technique in the future, but caution is advised

in the interpretation of analytical results from any external tissues due to the difficulty

of removing adhered contaminants. All Australian chelids have glands that secrete

musk (Legler & Georges 1993), the function of which is not known (Kool 1981), and

this secretion could also be tested for non-lethal contaminant biomonitoring. Another

method of simply determining whether or not turtles are currently exposed to

anthropogenic metals such as Pb, is by measuring faecal metal concentrations, which

may reflect variations in dietary intake (birds, Eeva & Lehikoinen 1996).

E4.4 Metals in Organs

Prior to the 1980s, inaccurate techniques meant tissue metal concentrations were

regularly overestimated (generally by 10-100 times, Friberg et al. 1979), but metal

assessment in reptiles was extremely rare, with most xenobiotic studies involving

organochlorine pesticides (reviewed in Hall 1980). Work on metal contaminants in

turtles began in the early 1980s with studies on metals in sea turtle eggs (Stoneburner

et al. 1980) and humeri (Witkowski & Frazier 1982), and assessment of Pb in the

bone & soft tissues of box turtles (Beresford et al. 1981). Tissue metal distribution in

reptiles, particularly at contaminated sites, is largely the same as for other vertebrates

(Linder & Grillitsch 2000). Environmental metal toxicity is most commonly

associated with the non-essential metals Pb, Al, and Ni (as well as Cd, As, Sn, Hg,

and Ag which were not detected in Sydney turtles), and the essential metals Cr, Cu,

Se, Zn, and Fe (as well as Mo which was only found near detection limits in liver and

kidney, and Co which was not detected in turtles) (Freedman 1995).

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E4.4.1 Liver & Kidney

The metal-binding metallothionein proteins, present in most eukaryotes including

freshwater turtles (Yamamura & Suzuki 1984; Trachemys scripta, Thomas et al.

1994), are present in all body tissues, but are usually in highest concentrations in the

liver and kidney (Linder & Grillitsch 2000), where they act as nontoxic metal

reservoirs under normal conditions (Hamer 1986, Sanders et al. 1996). However, their

high capacity for metal accumulation makes liver and kidney vulnerable target organs

for metal toxicity during environmentally-elevated metal exposures, when metal-

binding systems may become overloaded and excess metals start to interfere with

metabolic pathways (Squibb 1996).

Liver and kidney metal concentrations may vary with factors not related to

environmental loads. Chelodina longicollis were captured from the first half of

summer to the first half of autumn (Section D3.1.2), but due to the limited sample size

the effects of season on metal concentrations in these organs could not be analysed.

For future studies it should be noted that season may influence liver metal

concentrations (trends may differ over metals, Eastwood & Couture 2002). The effect

of sex on total organ metal burden should also be assessed, as in some species females

have significantly larger livers than males (at all times of year, Fletcher & King

1978), so that even with the same metal concentration females may carry the higher

load. Because of the capacity of liver and kidney for high levels of non-toxic metal

storage (Linder & Grillitsch 2000), the buffering capacity of these organs for elevated

intakes of environmental metals should be assessed in freshwater turtles.

Studies variably present metal concentrations as ‘per wet tissue weight’, ‘per dry

tissue weight’ (this study), or ‘per ashed weight’. For comparison with other studies,

per dry weight liver and kidney concentrations presented in this study are converted to

per wet weight by dividing by 4 (‘converted values’, Section E3.4). Unless otherwise

noted, all values in Chapter E4 are per dry weight.

Non-Essential Metals

Lead detection in some liver samples at up to 2.1 (average 0.65) mg/kg, and in one

kidney at 1.4 mg/kg contrasts to green turtles near the Hawaiian islands, where no Pb

was detected in any liver or kidney samples (Chelonia mydas, Aguirre et al. 1994),

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and to box turtles where Pb was found in all liver and kidney samples (Terrapene

carolina, Beresford et al. 1981). Pb concentrations in box turtle are similar in liver

and kidney (as for C. longicollis in this study), and both organs indicate the presence

of environmental Pb pollution. Pb concentration (mg/kg wet weight) averaged 21.6 in

liver and 24.3 in kidney for box turtles caught near Pb smelters, and 1.2 in liver and

1.8 in kidney for turtles caught distant to the Pb smelters (Beresford et al. 1981).

Converted values for C. longicollis liver (average 0.14, max 0.52 mg/kg wet weight)

are considerably less than averages for both box turtle populations. The average dry

weight liver Pb concentration (mg/kg) for C. longicollis (0.57) is higher than that

found for green turtles (0.51), hawksbill turtles (Eretmochelys imbricata, 0.17, Anan

et al. 2001), and a leatherback turtle (Dermochelys coriacea, 0.12, Davenport &

Wrench 1990); the difference is small considering the high Pb pollution found in the

Sydney environment (EPA 1996). Also, higher concentrations occur in other sea

turtles (1.2, Ca. caretta, Storelli et al. 1998). As Pb does not biomagnify in aquatic

food chains (Ward et al. 1986, reviewed in Prosi 1989), C. longicollis is not

disadvantaged by its high-order carnivory. Freshwater chelid turtles may have some

protective mechanism against Pb accumulation as, even in highly metal-polluted

environments, Pb was not detected in some liver and kidney samples of C. longicollis.

Titanium is the seventh most common metal, on average comprising 0.4% of the

earth’s crust (Moeller et al. 1984), but reaching 10% in some soils (reviewed in Berlin

& Nordman 1979). Ti is not known to be an essential element for any animal, and

because of its poor retention in animal tissues (Underwood 1977), and its inertness, Ti

and its compounds are widely used in surgical implants and as food colourants (Berlin

& Nordman 1979). Natural environmental Ti compounds tend to be poorly absorbed,

with inhalation of Ti-containing dust particles thought to be the cause of the highest

tissue concentrations being found in lung (humans, Berlin & Nordman 1979).

Ti has not previously been reported from turtle tissues, and has not commonly been

measured in any animal studies. In this study, Ti was found at a similar concentration

range in liver (0.8-2.1 mg/kg) and kidney (0.9-2.2 mg/kg). The 2.5x range from

minimum to maximum concentration is similar to, or less than, that found for

essential elements (Table E3.5). From this finding, and the presence of high Ti

concentrations in the carapace (Section E4.5), it is suggested that Ti may be an

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essential, or at least a utilised and possibly homeostatically-controlled, metal in C.

longicollis. The Ti concentration in liver is higher than that found in mammalian

aquatic vertebrates (<0.01-0.2 mg/kg, n = 3), although kidney concentrations may be

similar (0.3, 1.2 mg/kg, n = 2) (sperm whales, Holsbeek et al. 1999).

Essential Metals

The detection of Mn, Cu, and Se in all C. longicollis liver and kidney samples, and of

Cr in a proportion of liver and kidney samples, is largely consistent with that

previously found for sea turtles (Chelonia mydas, concentrations per wet weight,

Aguirre et al. 1994). Maximum Mn, Se, and Cr concentrations are similar in the two

organs for C. longicollis, whereas Cu was 5x higher in liver than kidney. Converted

maximum values for C. longicollis are lower than maximum values found for sea

turtles, except for Mn concentrations which are 2.2x higher in C. longicollis liver and

2.7x higher in C. longicollis kidney. This may be due to innate physiological

differences, or may be due to different environmental exposures.

Chromium is an essential element in animals (required for normal glucose

metabolism), is found in all organs but with highest concentrations in the lungs (from

particle inhalation), and is also concentrated in the liver (humans, Langård & Norseth

1979, Gerhardsson & Skerfving 1996). Cr enters the environment via fertilisers,

industrial and sewage wastes, and atmospheric dusts (Outridge & Scheuhammer

1993). Hexavalent Cr is the form that is of most toxic concern during environmental

exposure, with possible liver and kidney damage resulting from exposure (Ware

1988b). Most forms of Cr are reduced to and stored as the trivalent form in animal

tissue (Langård & Norseth 1979, Keen 1996), making it difficult to gauge toxicity

from tissue concentrations, although trivalent Cr predominantly deposits in the lungs,

whereas hexavalent Cr is more widespread, with deposition in kidneys, and increased

deposition in bone during chronic exposure (reviewed in Outridge & Scheuhammer

1993). Stainless steel contains about 18% Cr (Moeller et al. 1984), so the possibility

of Cr contamination of samples from scalpel blades should be acknowledged,

although with no Cr detected in some liver and kidney samples and all claw samples,

it appears that precautions against contamination were adequate. Also, stainless steel

contains about 8% Ni (Moeller et al. 1984), but this metal was not detected in any

liver or kidney samples.

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In aquatic systems, Cr is transferred through food chains but does not biomagnify

(Suedel et al. 1994), although bioconcentration factors (tissue/water concentration)

may reach 1000 (ANZECC 1992). In freshwater food chains, much higher

concentrations of Cr generally occur in invertebrates (e.g. annelids with 10 mg/kg wet

weight) than vertebrates (Mathis & Cummings 1973), although this may not apply in

urban areas as liver from frogs (Limnodynastes peronii) caught at a contaminated

pond in Sydney contained up to 185 mg/kg Cr (wet weight, Rader 1998). Average

concentration of Cr in liver (mg/kg dry weight) is lower in C. longicollis (0.31) than

in sea turtles (Ch. mydas 2.2, E. imbricata 0.85, Anan et al. 2001; Ca. caretta 1.1,

Storelli et al. 1998), and average concentration of Cr in kidney (mg/kg dry weight) is

lower in C. longicollis (0.19) than in sea turtles (Ch. mydas 2.2, E. imbricata 1.6,

Anan et al. 2001; Ca. caretta 1.6, Storelli et al. 1998). This may represent differences

in physiological requirements dependent on marine or freshwater environments.

However, average Cr concentration in these organs (C. longicollis values converted to

mg/kg wet weight: liver 0.08, kidney 0.05) is also higher for another freshwater turtle

(Chelydra serpentina, mg/kg wet weight) living in both brackish (liver 0.6, kidney

2.90) and freshwater (liver 1.03, kidney 1.10) (Albers et al. 1986; weighted mean

values recalculated, pooling sexes and sites). Values are also higher for T. scripta

from both metal-polluted (6.2 mg/kg dry weight) and non-polluted (1.2 mg/kg dry

weight) sites. The lower organ concentrations in C. longicollis suggest no cause for

concern over Cr toxicity in Sydney C. longicollis, and may simply reflect differing

physiological requirements (possibly at the pleurodire vs cryptodire level).

Manganese is an essential element for bone and connective tissue formation, and for

the metabolism of carbohydrate and lipids, and its absorption is homeostatically

controlled (Piscator 1979b, Keen 1996, Gerhardsson & Skerfving 1996). Ingested Mn

is relatively non-toxic, and the liver has a high capacity for Mn metabolism and

elimination (Pacyna 1996). Mn is distributed widely throughout tissues with highest

concentrations found in liver, kidney, endocrine glands, and bones (Underwood 1977,

Piscator 1979b, Gerhardsson & Skerfving 1996). The major impact of Mn deficiency

is on the bones (e.g. bone shortening, Keen 1996).

Liver usually contains 6-8 mg/kg Mn dry weight (humans, reviewed in Underwood

1977), which is within the range found in C. longicollis liver (4-24 mg/kg), indicating

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that turtles may have a lower degree of homeostatic control of Mn concentrations than

mammals. The average Mn concentration (mg/kg) in both liver (9.4) and kidney (5.2)

was similar to that found for sea turtles (Ch. mydas, E. imbricata; liver 4.7, 8.3;

kidney 5.6, 13.2; Anan et al. 2001). Although Mn toxicity is not apparent in C.

longicollis, Mn deficiency is indicated in turtles captured from the Sludge Lagoon

(Section E4.5.1)

Copper is an essential trace metal for animals (Keen 1996), being a cofactor in

enzymes associated with oxidative metabolism (Cohen et al. 1996). Liver and kidney,

along with brain and heart have the highest concentrations of Cu (reviewed in

Underwood 1977, reviewed in Piscator 1979a), although liver concentration may

decrease with age (Chelonia mydas, Sakai et al. 2000a). Temporary liver and kidney

damage may result after acute Cu ingestion (Piscator 1979a), and a Cu deficiency may

cause abnormalities in brain, bone and connective tissue (Keen 1996). For all species,

individual variation in liver Cu concentration is high, and concentrations may also

vary with species, age, diet, and disease, but not sex (Underwood 1977).

Cu does not biomagnify in aquatic systems (Mathis & Cummings 1973), but it may

bioconcentrate from 100-26000x (ANZECC 1992). The average liver Cu

concentration for C. longicollis (converted value of 9 mg/kg wet weight) is less than

that previously found for the genus (34 mg/kg wet weight, C. oblonga, n = 1, Beck

1956), although sample preparation and analytical technique differed greatly, and the

previous value is within the range for C. longicollis. The Cu liver concentration of C.

longicollis is also less than that for green sea turtles (50 mg/kg, Sakai et al. 2000a),

but it is very similar to that for loggerhead (Caretta caretta) and leatherback (D.

coriacea) sea turtles (both < 10 mg/kg wet weight, Caurant et al. 1999). The liver Cu

concentration range for C. longicollis (18-50 mg/kg dry weight) falls within the

common range of 10-50 mg/kg for reptiles and other vertebrates, although some have

a normal range of 100-400 mg/kg (dry weight, reviewed in Underwood 1977). An

average value in the high range group was found for Ch. mydas (139 mg/kg), while a

value above the common range was found for syntopic E. imbricata 55 mg/kg) (Anan

et al. 2001). In contrast, the average values in kidney for these species (Ch. mydas 8.3

mg/kg, E. imbricata 7.0 mg/kg, Anan et al. 2001) are within the range found for C.

longicollis (4-10 mg/kg, average 7 mg/kg). The concentration of Cu in the kidney of

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loggerhead (Ca. caretta) and leatherback (D. coriacea) sea turtles is also extremely

similar (2.2-2.7 mg/kg wet weight) (Caurant et al. 1999). From these comparisons,

there should be no concern over Cu toxicity in Sydney C. longicollis.

Selenium is an essential element for animals (Keen 1996). With little homeostatic

control of uptake, absorption of ingested Se can approach 100%, after which it is

distributed over all tissues, with highest concentrations in the liver and kidneys

(Glover et al. 1979). Se concentrations in these tissues can reflect a wide variation of

Se concentration in the food chain (Underwood 1977), including for reptiles (the

snake Pituophis catenifer, Ohlendorf et al. 1988). There is no evidence for the

biomagnification of Se in aquatic food chains (Suedel et al. 1994). However, liver Se

concentrations in birds can remain above concentrations associated with impaired

reproduction nine years after input of Se-contaminated water into wetland areas

ceases (Paveglio et al. 1997).

Chronic Se deficiency can cause disease of the liver, muscles, and pancreas (Keen

1996), whereas chronically elevated dietary Se results in liver damage, anaemia,

stunted growth, premature death, and adverse reproductive effects (Glover et al.

1979). Se toxicity causes birth defects in birds and fish, but not mammals, with birds

and fish sensitive to even small increases in environmental concentrations (reviewed

in Keen 1996). In aquatic birds, Se toxicity results in eye, limb, heart and gut defects

(reviewed in Keen 1996). Se also affects the functioning of almost all components of

the immune system (nonspecific, humoral, and cell-mediated), generally with Se

deficiency and Se toxicity both leading to immunosuppression, and low doses of Se

usually causing immunostimulation (Exon et al. 1996, reviewed in Sharma &

Dugyala 1996).

The livers of terrestrial animals usually have less than 1.5 mg/kg wet weight of Se

(Aguirre et al. 1994), which was the case for both liver (max 1.4) and kidney (max

1.3) converted values (mg/kg wet weight) in C. longicollis. This result for C.

longicollis, however, contrasts to that found for sea turtles, which may naturally have

higher organ Se concentrations, as they, generally without other elevated metal

concentrations, had wet weight liver Se concentrations generally considered toxic (2.5

mg/kg for Chelonia mydas and 3.4 mg/kg for C. caretta, Aguirre et al. 1994; 4.9

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mg/kg for C. caretta, Storelli et al. 1998). Some authors (Anan et al. 2001) have

found even higher concentrations (mg/kg dry weight) in sea turtles (liver 5.1 for C.

mydas, 49 for E. imbricata; kidney 5.3 for C. mydas, 28 for E. imbricata).

Accumulation of Se in marine animals is highly variable (0.05-30 mg/kg wet weight),

as evidenced by the much lower Se concentrations in D. coriacea liver (1.4 mg/kg dry

weight, n = 1, Davenport & Wrench 1990), and liver from Ch. mydas stranded on the

southeastern coast of Queensland (average 1.2 mg/kg wet weight, Gordon et al.

1998). Kidney Se concentrations in marine birds reach 10.2 mg/kg wet weight, a level

which is associated with impaired reproduction in shore birds (reviewed in Aguirre et

al. 1994). Normal Se concentrations in C. longicollis may in fact correlate better with

terrestrial animals than sea turtles. The difference between adequate and toxic Se

intakes can be small (Aguirre et al. 1994), so it is essential that baseline values and

toxic concentrations be established for Australian chelids. However it may eventuate

that this is not the case for freshwater turtles, as average Se liver concentrations (dry

weight) are higher in T. scripta from non-polluted sites (3.4 mg/kg), and much higher

in T. scripta from metal-polluted sites (37 mg/kg), without any overt consequences

(Nagle et al. 2001). Currently there appears to be no concern regarding Se toxicity in

Sydney C. longicollis.

Zinc and Iron were detected in all liver and kidney samples. Zn and Fe affect one

another’s absorption and storage (Underwood 1977, reviewed in Elinder & Piscator

1979a), and so should be assessed together in cases of toxicity or deficiency. As major

essential elements, these metals are subject to a high degree of homeostatic control,

and can be stored safely in tissues at relatively high concentrations as metal-protein

complexes (Elinder & Piscator 1979a, Sanders et al. 1996). High Zn concentrations

are found throughout the body, but especially in bone, muscle and liver (Elinder &

Piscator 1979b). Zn is efficiently excreted from the body (Sharma & Dugyala 1996),

does not biomagnify in aquatic systems (Ward et al. 1986), and has been historically

considered as nontoxic compared to the other metals (Elinder & Piscator 1979b).

However, a high Zn intake does cause secondary Cu deficiency (Keen 1996), and

there is some recent indication of direct toxic effects (Cohen et al. 1996).

Nonetheless, Zn deficiency is still considered much more problematic than Zn toxicity

(Cohen et al. 1996). It should be noted that during Zn deficiency, the organ Zn

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concentration may not be significantly lowered even though total Zn is reduced, as

organs may be smaller in Zn-deficient animals (Underwood 1977).

Average liver Zn concentration (mg/kg) for C. longicollis (dry weight 74; converted

wet weight 19) is less than that found in sea turtles (dry weight: 87 for Ch. mydas, 109

for E. imbricata, Anan et al. 2001; wet weights: 25 for Ca. caretta, 29 for D.

coriacea, Caurant et al. 1999; 40 for Ch. mydas, Gordon et al. 1998). Also, average

kidney Zn concentration (mg/kg) for C. longicollis (dry weight 59; converted wet

weight 15) is less than that found in sea turtles (dry weights: 169 for Ch. mydas, 120

for E. imbricata, Anan et al. 2001; wet weights 24 for Ca. caretta, 26 for D. coriacea,

Caurant et al. 1999; 21 for Ch. mydas, Gordon et al. 1998). Average values for three

Chelydra serpentina populations for liver are also higher than for C. longicollis (28.4-

47.3), but kidney values are lower (9.2-10.5, both wet weight; Albers et al. 1986),

suggesting that Zn concentrations in kidney may be generally lower in freshwater

turtles than sea turtles. The low range liver concentrations in C. longicollis suggest

there is no concern over Zn toxicity, and its ready environmental availability (Table

E3.3) suggests Zn deficiency is unlikely unless there are bioavailability issues.

Liver and kidney have a large capacity for non-toxic Fe storage and normal variation

in Fe concentration can be high within and between individuals, although these

organs, along with the spleen, usually have the highest Fe concentrations in the body

(Underwood 1977). However, as most absorbed Fe is stored not eliminated, chronic

intake of excessive quantities of Fe can cause damage to the liver (Clarkson 1979,

Elinder & Piscator 1979a). As with Zn, Fe deficiency is much more common and

problematic than Fe toxicity (Elinder & Piscator 1979a), with anaemia being a

common symptom (Keen 1996). Body Fe concentrations do not show much variation

in adult mammals (Underwood 1977), with the normal Fe concentration of human

liver being 500-800 mg/kg wet weight (reviewed in Elinder & Piscator 1979a). This is

similar to the converted concentration in C. longicollis liver of 185-784 mg/kg wet

weight, indicating that Fe deficiency is unlikely to be a problem in this species.

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Future Directions

The potential for non-toxic metal storage by metallothionein and other proteins in

turtle soft tissues needs to be ascertained so that tissue concentrations can be

interpreted in a more relevant manner. The fraction of the whole tissue metal

concentration that consists of free and available intracellular ions can be assessed for

some metals with the use of the fluorescent dye fura-2 (Biden & Browne 1993).

Measurement of metallothionein levels has been proposed as an indicator of metal

exposure and toxicity, although before this technique can be utilised, other factors

potentially inducing a stress response involving metallothionein production need to be

identified (Roesijadi 1992). Other influences on metal tolerance to be examined

include blocking metal entry into cells, extracellular scavenging of metals,

intracellular storage of metals in vacuoles, transport of metals out of cells, conversion

of metals into inert forms, and systems for repair of metal damage (reviewed in

Nieboer & Fletcher 1996). Understanding how metals cross cell membranes is, in

particular, fundamental to understanding metal toxicity (Chang 1996a). Neutron

activation and X-ray fluorescence are newer techniques which may be used for non-

lethal in vivo assessment of metals in organs (e.g. Cd in the kidney and liver and Pb in

bone (reviewed in Gerhardsson & Skerfving 1996), and will aid in the study of long-

term toxic metal effects (Hu 1998). For some metals, organ distribution can be

determined by following administration of radiolabelled ions, which could also help

determine if organ distribution is altered once the main target organs are overloaded,

and at what load this occurs.

E4.4.2 Bone

Bone is a major storage tissue for the alkaline earth metals (Be, Mg, Ca, Sr, Ba, Ra),

as well Pb (Camner et al. 1979). Also present are Mg, Zn, Cu, and Mn, which are all

required for normal growth and remodelling of bone (Bhattacharyya et al. 1996). The

non-essential elements (Pb, Ti, Ni, Sr, Ba Table E3.5; Al, Table E3.19) are mainly

deposited in bony tissues in C. longicollis.

Femur

A previous study on metals in turtle long bones (humeri of sea turtles, ashed weight,

Witkowski & Frazier 1982) detected Pb, Mn, and Zn – as found in femur in this study

– as well as Cu and Fe. The sea turtle ashed weight concentrations are not directly

256

comparable to this study, but proportionally the sea turtle humeri showed 2x as much

Mn, 7x as much Zn, and 15x as much Pb as the C. longicollis femurs, which suggests

that either Pb exposure was higher in the sea turtles, or that C. longicollis in some way

restricts Pb deposition in the endoskeleton.

Within individuals, Pb concentrations in humerus and femur are similar, and in box

turtles (Terrapene carolina) long bone Pb concentrations are approximately 2.5x

higher than those in liver and kidney (Beresford et al. 1981). The average femur:liver

ratio is much higher for C. longicollis (8.8), which may reflect different accumulation

patterns in the two species, or may indicate that Pb exposure is continuing in T.

carolina but historical in C. longicollis. As with the internal organs, the converted

(Section E3.4) average value for C. longicollis femur Pb (3.7 mg/kg wet weight), is

similar to that found in long bones of box turtles caught distant to Pb smelters (3.8

mg/kg wet weight), but much lower than that for turtles caught near Pb smelters (58

mg/kg wet weight) (Beresford et al. 1981), indicating that concentrations are not

extreme.

Carapace vs Femur

There are distinct differences in the metal composition of C. longicollis carapace and

femur (Table E3.5), with Ti and Cr both detected in carapace (up to 69 and 5 mg/kg

respectively), but not in femur, and Ni and Se both detected in femur (up to 0.8 and

2.5 mg/kg respectively), but not in carapace. This absence of metal detection in only

one of either carapace or internal bone was not found for C. caretta or C. mydas (Fe,

Mn, Zn, Cu, Pb, Ni, Cd, Hg), although Ti, Cr, and Se were not tested (Sakai et al.

2000b). Most of the eight metals analysed in C. caretta showed similar concentrations

in the two tissues, except for Fe, which was at 4x the concentration in carapace

compared to internal bone (Sakai et al. 2000b). Fe was not detected in either tissue in

dead C. longicollis, but was detected in the carapace of other C. longicollis (Sections

E3.5.4, E4.5). The contribution of blood to detectable Fe in bone is not known. In

contrast, Zn appears to be remarkably uniform in concentration across the tissues

(Table E3.5, Sakai et al. 2000b). Sr also shows similar concentrations in the two

tissues in several turtle species (C. longicollis, Table E3.5; T. scripta, C. serpentina,

Meyers-Schöne et al. 1993).

257

Pb concentration is significantly correlated in femur and carapace in snapping turtles

(Chelydra serpentina), with Pb concentrations (mg/kg wet weight) at a non-polluted

site the same in carapace and femur (1.0); however, as environmental Pb pollution

increased over sites, the Pb concentration increased to a much greater degree in femur

(115), than in carapace (33) (Overmann & Krajicek 1995), again highlighting

metabolic differences, and possibly differences in proportions of trabecular and

cortical bone. Variations in these factors, and the forces driving them, need to be

examined to help explain the different deposition patterns found in femur and

carapace in C. longicollis. Differences between the two bony structures could relate to

their storage functions, or to their strength requirements. In C. longicollis both femur

and carapace have a high proportion of trabecular bone, except towards the marginal

edge of the carapace where largely compact bone is present (Figures E3.3a, E3.3b).

Turnover is thought to be very low in the bone near the marginal edge of the carapace

in some turtles (Gopherus agassizii, Wronski et al. 1992), and differing requirements

for remobilisation and associated turnover rates may be playing a role in metal

concentration variation between carapace and femur.

Turtle morphology (Section A.1.1) and bone histology (De Ricqlès 1976) have

changed remarkably little over the past 200 million years. The ability of both endo-

and exoskeleton to absorb high concentrations of toxic metals without apparent ill-

effect (Overmann & Krajicek 1995) may have rendered the Chelonia tolerant of a

broad range of environmental metal conditions, thus contributing to their persistence

over time.

Claw

A heterogeneity of tissues are included in the claw samples – keratin, bone, and flesh

– making only general comments possible. Most distinctive about claw was that it had

the highest concentrations of Se and Zn of any tissue, which is not surprising as these

metals are often associated with proteinaceous tissues (Glover et al. 1979, Keen

1996). The converted mean value of 223 mg/kg Zn per wet weight in claw is much

higher than the 10-50 mg/kg wet weight usually associated with flesh (Elinder &

Piscator 1979b), and is higher than the converted mean value of approximately 90

mg/kg wet weight found for C. longicollis bone (carapace and femur, Table E3.5),

suggesting that the claw sheath is particularly high in this metal.

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E4.5 Carapace Metals

With the broad natural variation of metal concentrations in freshwaters and sediments

(Geary 1981), the infiltration of anthropogenic metals into areas remote from human

activity (Pacyna 1996), local variations in physicochemical (Section E4.2.2) and food

chain (Underwood 1977), factors that affect metal bioavailability, and differences in

tissue distributions between species (Llacuna et al. 1995, Vaneeden & Schoonbee

1996, Section E4.4), it is difficult to establish natural background metal tissue

concentrations. For this study, it was anticipated that park sites would provide some

‘urban background’ levels, but variation in abiotic metal concentrations are not

partitioned neatly between urban and park sites. Analysis of museum specimens could

show the historical variation in carapace metal concentrations at different sites in

Sydney, and provide a better indication of the range of ‘natural’ background tissue

levels.

Generally, after Ca, the essential metals in largest quantities in bone are Mg and Zn,

followed by much smaller quantities of Cu and Mn (Bhattacharyya et al. 1996),

which, if the alkali metals are excluded, is the case for C. longicollis carapace, except

that the order of Cu, Mn is replaced by the order Fe, Mn, Cr, Cu (Section E3.5.4).

Elevated concentrations of metals in bone and other tissues may not be toxic while

stored in inactive forms (Section E4.4.1), but it is generally not known at what storage

concentrations metals start to display toxicity and how this differs over organs (Beijer

& Jernelöv 1979b).

E4.5.1 Variations in Carapace Metal Concentration

Size/Age

The longevity and growth rate of an animal may have particular importance in aquatic

systems due to the uptake of many metals directly from water, rather than their

biomagnification through food chains (Beijer & Jernelöv 1979b, Suedel et al. 1994).

Depuration and effects of both essential and non-essential metals can be influenced by

age of the exposed organism (Nordberg et al. 1979). Tissue metal concentrations of

vertebrates are likely to increase with age, especially in long-lived species, and

especially for metals with a long biological halftime (e.g. Pb and Zn), and those

whose concentrations are not homeostatically controlled (reviewed in Linder &

Grillitsch 2000). For instance, tissue concentrations of Hg (Meyers-Schöne et al.

259

1993), Cu, Pb, and Zn may increase with total length or weight in reptiles, especially

at polluted sites (Linder & Grillitsch 2000).

There was no significant change in any metal concentration in carapace with turtle

size for either C. longicollis or E. macquarii (Section E3.5.1), and hence Hypothesis 2

is not rejected (Section E1.2). However, apart from one large juvenile C. longicollis

(CL = 133 mm) and two large juvenile E. macquarii (CL = 126, 137 mm), all animals

are in the adult size range (CL ≥ 140 mm, Sections D3.1.3, C3.3.2). Sensitivity to

metal toxicity varies widely with stage of development in amphibians (Diana et al.

2001), and although turtles do not undergo comparable developmental changes, the

adverse effects of metals on development may still render juveniles particularly

sensitive to metal accumulation. Diet also shifts with age in turtles (Clark & Gibbons

1969, Hart 1983), and this could provide variation in exposure between individuals

living within the same system. Other tissues could be more sensitive indicators of

changes in metal load with age. Future work incorporating a broad juvenile size range

and a variety of tissues would give better indication of metal load changes with age

and size. With the aging of turtles still problematic (Section B.2.3), fresh approaches

are required, and, for example, the use of an ion microprobe to measure metal isotope

ratios in calcified tissues (Stern et al. 1999) could be explored as an aging technique.

Sex

In general, tissue metal concentration does not vary with sex in reptiles (Linder &

Grillitsch 2000), including aquatic reptiles (C. porosus, Jeffree et al. 2001). This was

the case for most metals in both C. longicollis and E. macquarii carapace, and hence

for most metals, Hypothesis 3 (Section E1.3) is not rejected.

However, C. longicollis females have significantly lower Fe concentrations in the

carapace than males (Section E3.5.2). The quantity of blood in the carapace sample

could influence this result. Five metals have a higher concentration in blood than

carapace (Table E3.20). For Cu, Se, and Na, there is less than a 6-fold concentration

difference, for K there is a 16-fold difference, but for Fe there is a 355-fold difference.

Thus, if males have a more vascularised marginal carapace structure this could

account for the difference in Fe concentration. Variations in carapace structure

260

between males and females, in addition to differences in trends in metal deposition

and remobilisation, need examination in the future.

For E. macquarii, gravid females had higher carapace concentrations of Mg and Ba

than males and non-gravid females, while females usually had higher concentrations

of Ti and Sr than males (Section E3.5.2), although these trends need to be confirmed

with larger sample sizes. The tendency towards higher concentrations of the non-

essential alkaline earth metals Sr and Ba in females, may simply reflect higher

accumulation of Ca (Jeffree et al. 1995), possibly as a store for egg formation. The

molecular ratio of these and other bone-accumulating metals to Ca should be

examined in tissues in the future. Whatever the reason for variations in their

concentrations, neither Ti (Section E4.4.1) or the alkaline earth metals Mg, Sr, and Ba

(Section E4.3) are of particular toxicological concern for adults or embryos.

Mobilisation of tissue metals for incorporation into eggs could lead to lower

concentrations of metals in reproductively active female reptiles compared to males

and immature females, although the indication so far is that maternal metal loads are

not significantly affected (Linder & Grillitsch 2000). The general trend found in

carapace in this study was the opposite, and sexual differences are not known for

metals in the bony tissue of another aquatic vertebrate (C. porosus, Twining et al.

1999). The four metals elevated in E. macquarii females or gravid females are all

found in both eggshell and egg contents (Table F3.4), and another possibility is that

females normally accumulate higher concentrations of relevant egg metals (and hence

their non-essential analogs) so that they are available for egg production (Kenyon et

al. 2001). This does not afford an explanation for higher concentrations of Mg and Ba

only being found in gravid females, unless females accumulate more metal if they will

be reproductively active in the coming season. Again, metals should be examined in

their proportion to Ca. Future analysis of the ratio of bone metals to bone Ca, when

compared to the ratio of egg metals to egg Ca may give a better indication of whether

carapace is mobilised for egg formation, and the extent to which females excrete body

metal burdens via eggs.

Ca remobilisation from shell bone may also occur during the buffering of lactic acid

produced during extended anoxia (Warburton & Jackson 1995, Jackson et al. 1999).

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Males and females captured during the breeding season have often recently emerged

from a prolonged subaquatic torpor (Cann 1998), and if carapace has been

remobilised extensively in this period it may swamp any effect of metal release by

females for egg formation.

Sexual differences in metal concentrations occur for soft tissues in other turtle species

(Albers et al. 1986, Meyers-Schöne et al. 1993, Rie & Callard 1997). The situation

for hard tissues such as bone may be complicated by their heterogeneous turnover

rates, and the fact that Ca is integral to the bone structure (Section E1.1.1,

Bhattacharyya et al. 1996). If bone remobilisation simply leads to less bone, then

changes to bone metal concentrations will probably only occur if subsequent bone

deposition has different metal concentrations to original remobilised bone (Twining et

al. 1999). If the bone structure remains, but just becomes less calcified (e.g. the

conversion of compact to trabecular bone in turtle shell, Zangerl 1969), then bone

metal composition will vary dependant on whether all metals are remobilised in

proportion to their original concentration. Thus, physiological differences between the

sexes associated with differences in carapacial bone remobilisation may be difficult to

detect.

Sexing of turtles is problematic, especially for juveniles, and non-lethal sexing

techniques for Australian turtles should be assessed. They include hormonal assay and

laparoscopy (Rostal et al. 1994, Gross et al. 1995), with laparoscopy also able to give

information on the animal’s maturity and the stage of follicular development in

females (reviewed in Limpus & Reed 1985). In relation to reproduction, radiography

(Gibbons & Green 1979) is a non-invasive method of monitoring the number and size,

and possibly even the density of shelled eggs within the female.

Species

Different species show different metal tissue distributions. For instance, of four

tissues tested in birds collected from the same polluted wetland, highest Cd occurred

in the kidneys (3.4 mg/kg) of the ibis, the bone (5.1 mg/kg) of the coot, and the blood

(4.7 mg/kg) of the cormorant (Vaneeden & Schoonbee 1996). In some polluted

environments a metal may become elevated in the bone of one species, but remain at

normal concentrations in another (birds, Llacuna et al. 1995).

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For most metals, there are no differences between the two species of turtles, and

hence Hypothesis 4 (Section E1.2) is not rejected. The few differences in the carapace

metal concentrations between C. longicollis and E. macquarii do not occur

consistently among sites, and further study with a larger sample size is required to

confirm them.

The differences between the two species in the concentration of the alkaline earth

metals (Ca, Sr, Ba) may be associated with a higher degree of calcification in the

marginal edge of the carapace in C. longicollis, as it is much thicker and harder than

in E. macquarii (pers. ob.). Zn and Ti are the only other metals showing any

indication of interspecific variation. Only Ti, which is proposed as a possibly essential

metal in C. longicollis (Section E4.4.1), showed a species difference at more than one

site. If the function of Ti is to add strength to calcified tissues, the higher

concentrations in C. longicollis may contribute to its harder shell.

For turtles with similar dietary differences to C. longicollis and E. macquarii, Hg has

previously been detected at higher concentrations in predominantly carnivorous

compared to predominantly herbivorous freshwater turtles (Meyers-Schöne et al.

1993). Hg is one of the very few metals with the potential to biomagnify in aquatic

systems (Suedel et al. 1994), and the species difference is probably due to

interspecific variation in diet, not physiology, as other metals do not differ between

carnivorous and omnivorous vertebrates (fish, reviewed in Pip & Stepaniuk 1997).

Metal concentrations in soft tissues may vary between male and female turtles

(Meyers-Schöne et al. 1993, Rie & Callard 1997), but within a species this difference

may occur at some sites but not others (Albers et al. 1986), and this complicating

factor needs to be resolved for Australian freshwater turtles for carapace, as well as

for internal tissues.

Metal absorption can range from less than 1% (trivalent Cr, Gerhardsson & Skerfving

1996) to 95% or more (Se, Glover et al. 1979; methylHg, Mason 1996), yet at present

there is virtually no information on the intestinal uptake of metals by reptiles (Linder

& Grillitsch 2000), and no comparisons of uptake between sympatric turtle species.

The gastrointestinal tract of turtles takes several days to empty (Moon & Foerster

2001), with the total gut retention time being affected by temperature and particle size

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(Spencer et al. 1998) and possibly affecting metal absorption (Twining et al. 1999).

The impact of gut retention time, particle size, temperature, pH, metal

physicochemical form, and dietary components (other metals, anions, chelating

agents) on the degree of metal uptake from ingested items and subsequent

assimilation into tissues (Camner et al. 1979, Nordberg et al. 1979, Pacyna 1996,

Linder & Grillitsch 2000) should be determined for freshwater turtles. Metal

absorption in turtles can be determined by measuring the percentage of a metal dose

that emerges from the gut in the faeces.

Compared to E. macquarii, which shows a reluctance to migrate even short distances

(Section A.1.4.1, C4.3.3), one might expect a greater variation in tissue metal

concentrations from C. longicollis because of their more migratory tendencies

(Section A.1.3.4) and consequent exposure to greater environmental and dietary metal

variations. However, the ranges of metal carapace concentrations are broader in the

majority of cases for E. macquarii than C. longicollis (Table E3.6), possibly

indicating less homeostatic control of uptake.

Sites & Urbanisation

Of the 16 carapace metals detected (Section E3.5), the concentrations of only two (Al

and Se) showed no difference over sites (Table E3.19). Five of 18 metals detected in

crocodile bone (C. porosus osteoderm) also differed over sites, with the association of

tissue metal concentrations with site of capture indicating some degree of site fidelity

(Jeffree et al. 2001). Differences across sites may reflect differing metal

bioavailabilities, including varying loads in dietary food chains. Thus, although

carapace metal concentrations do not reflect internal organ metal concentrations

(Section E4.3; Hypothesis 1 not rejected, Section E1.2), or current environmental

concentrations (Section E3.3.4; Hypothesis 6b not rejected, Section E1.2), carapace

analysis may still be a valuable technique for measuring historical metal exposure,

although the time period represented by bone samples is at present unknown. The five

(Mn, Co, Fe, Mg, U) of 18 metals detected in crocodile osteoderm that differed

between catchments were different from the five metals that differed in flesh (Pb, Al,

Ni, Cr, Ba) (Jeffree et al. 2001), so analysis of a soft tissue in conjunction with

carapace is indicated for biomonitoring purposes.

264

Saline water may lead to an increased tissue concentration of metals in snapping

turtles (Section E4.2.1). Of all nine sites, the saline Marley Lagoon had the highest

carapacial concentration of Ti, Ca, and, not surprisingly, Na (Tables E3.10-E3.18),

none of which are metals of toxicological concern. Marley Lagoon also had the lowest

concentration of Cu and K, and the second lowest concentration of Zn, suggesting that

there is no overall effect of water salinity on metal accumulation in carapacial tissue,

although a controlled experiment would be required to confirm this.

Mg, Zn, Cu, and Mn, are all involved in normal bone growth and remodelling, and

deficiencies lead to structural defects, as do elevated concentrations of osteotoxic

metals such as Cd, Pd, and Al (Bhattacharyya et al. 1996, Eeva & Lehikoinen 1996).

Turtles from the Sludge Lagoon had noticeably crumbly carapacial bone, yet

compared to other sites, these turtles have low concentrations of the osteotoxic metals,

Pb and Al, and similar concentrations of the essential Mg, Zn, and Cu. Although low

bone Cd concentrations (0.1 mg/kg dry weight) cause bone loss and increased bone

fragility (reviewed in Bhattacharyya et al. 1996), no Cd was detected in carapace in

this study. Sludge Lagoon turtles have the lowest average Mn concentration, the

lowest minimum Mn concentration, and, aside from U1, the lowest maximum Mn

concentration, so low Mn (Section E4.4.1) may be responsible for the crumbly bone

found at the Sludge Lagoon, although the underlying cause is unknown.

Ca tissue concentration explains 87-98% of variation in the tissue concentration of Ba

and Ra in freshwater bivalves (Jeffree et al. 1995), and this relationship should be

examined in the future to see if it explains variations in carapace metal concentrations

in individual turtles from the same site, especially for Pb, Sr, Ba, and other metals that

deposit with Ca. This may also help explain intersite differences as Ca is among the

metals whose tissue proportions are most associated with site of capture (Markich et

al. 2002).

E4.5.2 Non-Essential Elements – Specific Comments

Pb was detected in all but one carapace sample, with absorption likely to have been

through both pulmonary and gastrointestinal surfaces. Pb constitutes 1% of Sydney’s

fine atmospheric particles (Cohen et al. 1994, as cited by EPA 1996), and about 30%

of the inhaled Pb will be absorbed directly by the lungs, or after clearance to the

265

gastrointestinal tract (Tsuchiya 1979). There is no homeostatic control of Pb uptake,

so gastrointestinal uptake from ingested substances continues even after high tissue

Pb concentrations are reached (Ragan 1983). Although Pb uptake in urban areas is

therefore likely to be high, the majority of the Pb will be stored in bone, where it is

largely inert (Clarkson 1979), and Pb toxicity will generally only be realised when

bone is remobilised (Gerhardsson & Skerfving 1996). Mobilisation of Pb from

trabecular bone and soft tissue pools by administration of a chelating agent (e.g.

EDTA) may be a method of indicating the metabolically active, and therefore

toxicologically-relevant, Pb pools (reviewed in Gerhardsson & Skerfving 1996),

possibly providing a better indication of immediate toxic risks in turtles than

carapacial bone analysis.

A deficiency of an essential metal may cause an increase in the accumulation of one

or more non-essential metals, and this includes a Zn deficiency resulting in increased

Pb absorption (Keen 1996) and increased deposition of Pb in bone (Nieboer &

Fletcher 1996). What would constitute a Zn deficiency in Australian freshwater turtles

and how it would manifest in the carapace is unknown. However, although Zn

showed a two-fold change in average carapace concentration (74-142 mg/kg) among

sites, the 4 sites with the lowest Zn concentration (U3 < P4 < P3 = Sludge) did not

include any of the 4 sites with the highest Pb concentration (U1 > U2 > U4 > P1),

discounting this as an issue in this study.

In human bone, of 9 metals analysed (Pb, Al, Fe, Zn, Mg, Ca, Sr, Na, K), only the

concentration of Pb differed between osteoarthritic bone and normal bone, and it was

thought that the lower Pb concentration in osteoarthritic bone may reflect increased

bone turnover in this condition (femurs, Roberts et al. 1996). The Sludge Lagoon had

the lowest average carapacial Pb concentration (16 mg/kg) apart from one park site (8

mg/kg at P3), and also had distinctly fragile bone, suggesting there may be a link. All

other sites are similar in their Pb concentrations (21-28 mg/kg) except for U1 (68

mg/kg) and U2 (46 mg/kg). These elevated concentrations are similar to Pb

concentrations in the bone of urban pigeons (51 mg/kg), with the other sites having

similar concentrations to non-urban pigeons (30 mg/kg) (Janiga & Žemberyová

1998), suggesting U1 and U2 may be subject to aspects of urbanisation that are not

present at U3, U4, or the park sites. The elevated carapacial concentrations of Pb at

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U1 and U2 reflected the sites’ elevated water Pb concentrations (Section E4.2.2), as

also shown by crocodiles originating from Pb-contaminated areas (C. porosus

osteoderm, Twining et al. 1999). The carapacial Pb concentrations of snapping turtles

also reflects environmental contamination at capture sites with average concentrations

up to 33 mg/kg wet weight in highly disturbed and Pb-polluted areas (C. serpentina,

Overmann & Krajicek 1995). The turtles appear to live equally well in sites with high

and low Pb contamination, with no apparent effects on growth, and freshwater turtles

may generally survive well in areas polluted with metals (Nagle et al. 2001).

Individual carapace Pb concentrations over sites ranged from 0-110 mg/kg, which is

much greater than the ranges found in humans for both trabecular bone (0.07-7.1

mg/kg) and compact bone (0.25-22 mg/kg) (Drasch et al. 1997). In humans, bone Pb

varies significantly but consistently over the skeleton, so that if differences in the

proportion of trabecular to cortical bone are taken into account, analysis of Pb from a

single bone sample can be used to indicate the total skeletal burden (Hu 1998). Mice

near a heavily-trafficked road (5000 vehicles/h) have much higher bone Pb

concentrations (398 mg/kg) than mice near a low traffic road (10 vehicles/h) (87

mg/kg), with the high Pb group showing adverse reproductive effects (abnormal

sperm cells, Ieradi et al. 1996). Similarly, urban turtles are subject to a range of Pb

pollution. The point at which Pb toxicity impacts on reproduction in Australian

freshwater turtles will need to be resolved by laboratory experiment due to the

multiplicity of other factors at field sites.

Al toxicity is associated with impaired reproduction and reduced survivorship in

aquatic animals (Sparling & Lowe 1996). Al also inhibits bone mineralisation thus

increasing bone fragility (Mjöberg et al. 1997), although this does not appear to be

contributing to the fragile carapace bone of the turtles from the Sludge Lagoon, as

their Al concentrations are lower than at the majority of other sites. Al occurs in high

concentrations in lagoon sediments (Table E3.4) and preferentially accumulates in

bone in many birds and mammals (Bhattacharyya et al. 1996), but was only detected

in 1/3 of carapace samples (data not shown). Tissue distribution varies however, with

the highest concentrations of Al found in lung in humans (Norseth 1979), and Al was

detected in muscle (90 mg/kg) but not bone in crocodiles (Jeffree et al. 2001). Bone,

along with brain, are the critical organs for Al toxicity (Gerhardsson & Skerfving

267

1996), so if Al is also showing minimal accumulation in turtle brain, there is little

cause for concern over toxicity. Al is soluble and bioavailable at pH < 5.5, but not at

pH 5.5-7.5, with bioavailability at pH > 8 not well understood (Sparling & Lowe

1996). For birds it is certainly in acidified habitats that Al affects bone and eggs

(reviewed in Miles et al. 1993). The fact that all tested sites have pH > 5.5 (Table

E3.2) helps explain the lack of Al in carapace bone, despite high environmental

concentrations. Accuracy of Al detection was less than for most other metals (Section

E3.2.1) and values are often bordering detection limits, so detection of Al at over 1

g/kg in three samples (Section E3.5.4) should be interpreted with caution, although Al

concentrations may exceed 17 g/kg in aquatic invertebrates (Sparling & Lowe 1996).

Ti was detected in all carapace samples (up to 81 mg/kg), which contrasts to crocodile

(Crocodylus porosus) osteoderms where Ti was not detected, despite being present in

muscle (6.2 mg/kg, Jeffree et al. 2001; Markich et al. 2002). Neither was Ti detected

in C. longicollis femur (Table E3.5, Section E4.4.2). Ti forms strong lightweight

alloys, but is thought to be biologically inert (Berlin & Nordman 1979). Identification

of the chemical form of Ti in the carapace may shed light on whether it could be

intentionally sequestered for the increasing of shell strength or some other purpose. If

Ti is being taken up indiscriminately, for example through Ca pathways, one would

expect it to deposit in all bony tissues. Its very high concentration in carapace, yet

absence from femur, suggests its destiny is controlled independently to Ca.

Sr is elevated in the carapace of turtles (Trachemys scripta, Chelydra serpentina)

caught at contaminated sites (90Sr, Meyers-Schöne et al. 1993), yet in Sydney turtles,

average Sr concentrations in carapace are higher at all Park sites compared to all

Urban sites, despite the Urban sites having the majority of higher Sr concentrations in

both water and sediment (Table E3.3). Thus, although environmental Sr load may

play a role in Sr tissue accumulation, other factors must also be involved.

E4.6 Blood Metals

Whole blood analysis is a safe method of monitoring metals in live turtles (Kenyon et

al. 2001), with whole blood analysis preferential to serum or plasma analysis, because

metals adhere to erythrocytes, leukocytes and plasma proteins (Dessauer 1970). Blood

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metal analysis is a useful technique as blood integrates metal uptake by all body

surfaces, and can reflect concentrations in soft tissue pools including target organs, to

give a good indication of potential for toxic effects (Camner et al. 1979).

Environment/Species

There are no significant differences in blood metal concentrations between U3 C.

longicollis and U4 E. macquarii, except for Fe, which is higher in U4 E. macquarii

(Section E3.6.1). Although sediment Fe concentrations are also higher at U4 (Fe not

detected in lagoon waters), no conclusions can be drawn as to the relevance of this, as

Fe bioavailability and passage through the food chain is not known. Also, homeostatic

mechanisms tightly control Fe uptake from the environment (Elinder & Piscator

1979a), with extremely high intakes required to overwhelm this (Clarkson 1979).

Elevated Co, Cu, and Zn concentrations all decrease Fe absorption (reviewed in

Elinder & Piscator 1979a). In water, Co and Cu were not detected, and Zn

concentration was similar at both sites (Table E3.3), but in sediment the

concentrations of all three metals are at least 3x higher at U4 than at U3 (Table E3.4).

Thus, Fe accumulation is not decreased by high environmental Co, Cu, and Zn in

these turtles.

Another environmental factor that could affect Fe concentration in blood is the

dissolved oxygen concentration of lagoon water. Erythrocyte enlargement may be a

way of increasing blood oxygen-carrying capacity when aquatic oxygen

concentrations are low (Section D4.4 – Site), and this would result in an increase in

Fe concentration in the blood due to increased haemoglobin content. The Fe content

of haemoglobin is around 0.35% (Underwood 1977), with over 99% of Fe in blood

found within the erythrocytes bound to haemoglobin (Elinder & Piscator 1979a).

Surface oxygen concentration was actually higher on the day of testing at U4 than U3

(Table E3.2), but U4 is anoxic below one metre (Section E4.2.1), so low aquatic

oxygen concentration may be contributing to high blood Fe at this site.

The difference in Fe concentration may be dependent on species rather than on site. In

adult mammals there is little difference over species in total body Fe concentrations

(Underwood 1977), but the degree of variation is not known for reptiles. Age, sex,

diet, pregnancy, and disease state all also affect the quantity of haemoglobin, and

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therefore Fe, in the blood (Underwood 1977). For instance, green turtles with severe

fibropapillomatosis have low blood Fe concentrations (Aguirre & Balazs 2000).

Feeding, diving, temperature, and water availability can also greatly affect blood salt

concentrations (Dessauer 1970), and blood metal concentrations may also vary with

turtle size (Lepidochelys kempii, Kenyon et al. 2001). Further study is required to

resolve the degree of contribution from these factors in Australian chelids.

Sex

There are no significant differences between male and female C. longicollis at U3 in

any blood metal concentrations, and Hypothesis 3 (Section E1.2) is not rejected. An

absence of sexual difference was also found for blood metals (Pb, Hg, Ag, Cu, Zn) in

sea turtles (Lepidochelys kempii) (Kenyon et al. 2001), and further study will reveal

whether this is the case throughout the Chelonia. Sexual differences that are not

apparent in whole blood may become evident in plasma at certain times of year, with

Ca, Mg, Zn, and Fe in females increasing to significantly different levels from males

during vitellogenesis (alligator plasma, Lance et al. 1983).

Blood vs Carapace

There was no consistent pattern of correlation between blood and carapace metal

concentrations across sites and species, and hence Hypothesis 1 (Section E1.2) is not

rejected. However, there are some instances of correlations of metal concentrations in

carapace and blood within sites. Usually about 90% of Pb is stored in bone, where it

has a halftime of around 20 years, with the remainder found in blood, soft tissue, and

some high turnover bone pools, where it has a halftime of about 20 days (Tsuchiya

1979). Thus, blood Pb levels are generally well correlated with current Pb exposure,

but commonly do not correlate with bone Pb (Hu 1998). Metal concentrations in

blood and carapace did not correlate in U3 C. longicollis, but there was a correlation

in Pb concentrations between the two tissues in U4 E. macquarii (Section E3.6.2).

Thus there appears to be a correlation between current (blood) and historical (bone)

Pb pollution at U4. In addition to assimilation from recent environmental sources,

blood Pb also originates from the mobilisation of previously deposited bone stores

(Eisenmann & Miller 1996), which can cause considerable ‘endogenous’ Pb exposure

(reviewed in Gerhardsson & Skerfving 1996). All except one U4 E. macquarii are

females, so it is possible that bone remobilisation was taking place at this stage for

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egg formation and leading to an increase in blood Pb. This possibility is supported by

the fact that the one male blood Pb value was below the lower confidence limits for

female values (Table E3.21).

E4.6.1 Blood Metal Toxicity

Interpretation of blood metal concentrations in reptiles can be difficult due to the

sparsity of similar studies (Chiodini & Sundberg 1982, Kenyon et al. 2001), and

analyses being variably performed on whole blood, plasma, or serum. For comparison

with other studies, dry weight blood metal concentrations found for Sydney chelids

are divided by 7.7 for conversion to per wet weight concentrations (Section E3.6.1).

Of the eleven metals detected in the blood of C. longicollis and E. macquarii, five

(Mg, Ca, Sr / Na, K) are alkaline earth or alkali metals whose normal concentrations

(Sr is non-essential) in turtle blood are unknown.

Three of the four essential trace elements (Cu, Se, Zn, but not Fe) found in C.

longicollis and E. macquarii blood (converted to mg/kg wet weight) were also

measured in the blood of slider turtles (T. scripta, mg/kg wet weight) from three lakes

in Texas (Clark et al. 2000). For Cu, the averages for C. longicollis (0.61) and E.

macquarii (0.46), both lay within the range of averages for the three T. scripta

populations (0.42, 0.50, 0.74). The total Cu concentration range for the T. scripta

populations (0.13-1.3) also encompassed the ranges for both C. longicollis (0.51-0.74)

and E. macquarii (0.39-0.52). For E. macquarii, the average (0.24) and range (0.21-

0.28) for Se was again encompassed by values for the T. scripta populations. On the

other hand, C. longicollis showed a higher average (0.50), and a higher maximum

value (range 0.33-0.83) than T. scripta (population averages 0.17, 0.24, 0.26; total

range 0.11-0.51), although the ranges overlapped. Averages and maximum values for

Zn are slightly lower for both C. longicollis (average 4.4, range 3.5-5.3) and E.

macquarii (average 3.9, range 3.1-4.6) compared to the three T. scripta populations

(averages 4.9, 4.8, 5.5, total range 3.4-6.7), although all ranges overlap. Sea turtles

also intersect this Zn range, although average (7.5) and maximum (range 3.3-19)

values are higher (Lepidochelys kempii, Kenyon et al. 2001). The tendency to lower

blood Zn for the chelids and higher blood Se for C. longicollis was minor and not

suggestive of deficiency or toxicity. However, the full meaning of the blood metal

concentrations will not be clear until population variation and normal blood

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concentration ranges for these metals are established for the different species.

Although plasma metal concentrations are not measured in this study, the finding that

the normal plasma Zn concentration in turtles is 4-8 x, and snakes up to 50 x, that in

mammals (Lance et al. 1995), once again emphasizes the need for baseline data of

metal concentrations in different taxa.

Of the non-essential trace elements, Pb, in contrast to Ti, has received a lot of

attention due to its toxic effects in humans, and concentrations have been reported

from a wide range of tissues over a wide range of taxa. Blood Pb concentrations are

normally reported per wet weight or per volume of tissue. The converted (μg/kg wet

weight) mean Pb concentration for C. longicollis is 25 μg/kg (range 0-83) and for E.

macquarii is 114 μg/kg (range 55-190). These average and maximum values are

higher than those found in Kemp’s Ridley sea turtle blood (Lepidochelys kempii;

average 11, range 0-34 μg/kg) (Kenyon et al. 2001), suggesting a lower degree of

current exposure or uptake in the sea turtles. As with Pb in liver and kidney (Section

E4.4.1), blood Pb concentration was lower in C. longicollis than in box turtles (T.

carolina, Beresford et al. 1981). The average blood Pb for E. macquarii is similar to

that found for box turtles captured distant from Pb smelters (100 μg/kg), but much

lower than that for box turtles caught near to Pb smelters (6000 μg/kg, Beresford et al.

1981), suggesting current, but not extreme, pollution at the Sydney site. The Pb

concentration in box turtles was greater than that causing weakness and reduced

ability to walk in an adult snapping turtle (3600 μg/kg, C. serpentina) following

ingestion of a large Pb sinker and fishing hooks (Borkowski 1997), indicating that Pb

tolerance varies between species. Blood Pb in the snapping turtle dropped to

undetectable levels following chelation therapy (Borkowski 1997), with the turtle’s

recovery indicating an ability in the Chelonia to recover from extreme temporary

elevated Pb exposures. At the same dietary Pb intake, low dietary Ca results in greatly

increased Pb concentrations in blood (Six & Goyer 1970), so the high Ca levels at the

higher environmental Pb sites in Sydney may be having a protective effect for blood,

as was suggested for bone.

Once blood Pb reaches 100-150 μg/l there may be impaired neurological function in

rodents and human juveniles, and once levels over 400 μg/l are reached all body

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systems, including the reproductive system, are at risk of adverse effects (reviewed in

Audesirk & Audesirk 1996, and Eisenmann & Miller 1996). The snapping turtle, and

the high Pb population of box turtles, have blood Pb concentrations elevated greatly

above these levels, indicating an astounding tolerance to Pb toxicity. In other turtles

(T. scripta) an injection of 1 mg Pb/g body weight into hatchlings does not alter

survival or behaviour (Burger et al. 1998). In Sydney, only some E. macquarii are

above the level associated with impaired neurological function, and no animals have

blood Pb over 400 μg/kg, so there is little cause for concern over Pb causing

reproductive problems in these populations, although behavioural changes could

occur.

In humans, high blood Pb occurs in populations living near heavily trafficked roads

and in other urban areas, with a subsequent decrease in blood Pb levels from urban to

suburban to rural populations (reviewed in Tsuchiya 1979). Prior to the introduction

of unleaded petrol, blood Pb (‘μg%’ converted to μg/l wet weight by multiplication by

10) in children at a school near U2 was 250 μg/l (Gardeners Rd, range 70-600),

whereas at a school just outside the Sydney Basin (Blaxland East) blood Pb in school

children averaged 160 μg/l (range 70-400) (Garmys et al. 1979). The significant

difference between these and two other groups of school children was directly related

to variations in atmospheric Pb concentrations, and turtles within and away from the

Sydney Basin are likely to be subject to similar exposure gradients. Increases in blood

Pb correlated with increases in behavioural problems in the school children (Garmys

et al. 1979), and loss of the righting response in turtles is directly related to increasing

Pb dose (Burger et al. 1998). Australian chelids show distinct mating behaviours that

are somewhat more elaborate in the shortnecked genera (Murphy & Lamoreaux 1978,

Banks 1987, Georges et al. 1993). It is possible that Pb, which in humans causes

subtle behavioural changes before having any more noticable effects, has an effect

courtship behaviour and therefore on copulation success. Aside from reproductive

consequences, metal toxicity can affect behavioural functions such as searching for

food, recognising predators, and proper orientation (Zillioux et al. 1993). All these

behaviours have the potential to be adversely affected by Pb toxicity, although at this

stage it appears that behavioural changes may only occur at similar, or slightly lower,

concentrations than those that cause death in turtles (T. scripta, Burger et al. 1998).

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During drought, Chelodina spp. may aestivate on land by burrowing under soil or

litter, or by encasing themselves in the sediments of drying water bodies (Chessman

1983, Ehmann 1992). Aestivation may result in a drop of almost a third to the turtle’s

metalbolic rate (Kennett & Christian 1994), and aestivation chambers may also reduce

daily temperature fluctuations by over 30 ºC and maximum temperatures by 18 ºC

(compared to soil surface, Kennett & Christian 1994). No one has considered what

changes metabolic depression or reduced body temperature will have on metal

physiology and toxicity. Analysis of blood metals during this time may indicate

whether aestivation has a protective effect. Also, C. longicollis only grow for six

months of the year (Kennett & Georges 1990), and the effect of pollutant exposure to

turtle health or reproduction may be altered during periods of assimilation and

growth, compared periods of dormancy. Again, blood analysis is probably the most

efficient way of detecting potential differences.

E4.7 Summary

There are no studies published on the effects of bioaccumulating pollutants on

freshwater turtles in Australia, and very little on any other Australian reptile (some on

sea turtles and crocodiles). The few studies conducted on metal contaminants in

turtles have involved sea turtles or the freshwater turtles of North America.

Freshwater turtles are considered excellent potential bioindicators (Meyers-Schöne &

Walton 1994), in part because they are often long-lived dietary generalists that are in

close association with aquatic and sediment metals (Sparling et al. 2000a), and also

because of their persistence in contaminated environments (reviewed in Albers et al.

1986). The potential for C. longicollis and E. macquarii as sentinel species for metal

accumulation was examined. Metal concentrations were analysed in five tissues from

dead turtles (liver, kidney, femur, carapace and claw) and fifteen metals (Pb, Ti, Ni,

Cr, Mn, Cu, Se, Zn, Fe, Mg, Ca, Sr, Ba, Na, K) were detected in one or more tissues.

Correlations in metal concentrations largely occurred for bone-accumulating metals

among the bone-containing tissues. Metal concentrations in the two tissues that could

be sampled non-lethally (carapace and claw) did not reflect concentrations in internal

organs, except for a few exceptions that did not include any metals of great

ecotoxicological relevance. Thus, carapace analysis is largely discounted as a

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potential tool for monitoring metal loads in internal chelid organs (Hypothesis 1,

Section E1.2).

Metal concentrations in C. longicollis and E. macquarii carapace were measured in

free-ranging turtles captured at four urban sites (presumed to be highly polluted due to

their proximity to industrial or dense urban development), and from four nearby sites

in national parks (presumed to be minimally polluted). Carapace metals from turtles

from the sludge lagoon from Section D were also analysed. Only two urban sites had

water metal concentrations at levels of concern (Pb, Cu, Zn, Fe). The 29 metals found

in sediment varied in concentrations between sites, with individual urban sites often

having the higher concentrations, but without any overall significant differences

between urban and non-urban sites (Hypthesis 5, Section E1.2).

Sixteen metals (Pb, Al, Ti , Zr, Cr, Mn, Cu, Se, Zn, Fe, Mg, Ca, Sr, Ba, Na, K) were

detected in some or all carapace samples from these live turtles. Significant

differences are present between sites for most carapace metal concentrations, but the

division was not between urban and non-urban sites. Even though carapace was not

thought useful for the indication of internal metal burdens, the site differences

indicate the potential of carapace metal concentrations to reflect long-term metal

presence in different environments. There are few significant differences in carapace

metal concentrations between C. longicollis and E. macquarii and these are not

consistent over sites (Hypothesis 4 not rejected, Section E1.2).

Metal concentrations in carapace did not change with increasing turtle size (although

few juveniles are available) (Hypothesis 2 not rejected, Section E1.2), and for most

metals there are no differences between the sexes (Hypothesis 3 not rejected, Section

E1.2), or between gravid and non-gravid turtles (with the exception of Fe in C.

longicollis, and Ti and Sr in E. macquarii). Lower metal concentrations may be found

in the bony tissues of gravid females compared to juveniles or males due to the

remobilisation of Ca and other metals for egg formation. There is a significant

negative effect of maternal carapace concentrations of Ca and Mg on eggshell

thickness in E. macquarii, indicating that there may be mobilisation of Ca and Mg

from the carapace for eggshell formation. The maternal tissue or tissues from which

other egg metals originate remains obscure however, as femur differs from egg tissues

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by containing Pb, but not Ti, Cu, or Fe, and carapace differs from egg tissues by

containing Pb, Al, Zr, and Cr, but not Ni. Egg metals may also be originating from

other maternal stores, or from current environmental uptakes.

Metal concentrations in blood were measured in C. longicollis at one urban site, and

E. macquarii at another urban site. Only Fe concentration differed between the

species/sites, with the higher concentrations in E. macquarii possibly due to an

increase in haemoglobin resulting from the site’s low aquatic oxygen concentration.

Carapace metal concentrations are not correlated with water or sediment metal

concentrations, except for the correlation between aquatic Pb and carapace Pb

(Hypothesis 6 rejected for Pb only, Section E1.2). Also, although there are no

correlations between blood and carapace metal concentrations for most of the metals

found in chelids, there was a correlation between blood and carapace Pb

concentrations in E. macquarii. Blood may be a useful biomonitoring tissue for

detecting body Pb loads resulting during current exposure to bioavailable

environmental Pb.

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Section C: HM

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